The Response of Aquatic Macrophytes to Lake Management Practices and the Role

of Light in the Germination of Macrophyte Propagules

A THESIS

SUBMITTED TO THE FACULTY OF

UNIVERSITY OF MINNESOTA

BY

MELANEY A. DUNNE

IN PARTIAL FULFILLMENT OF THE REQUIREMENTS

FOR THE DEGREE OF

MASTER OF SCIENCE

ADVISOR: DR. RAYMOND M. NEWMAN

September 2017

© Melaney A. Dunne 2017

Acknowledgements

Firstly, I thank my advisor, Dr. Ray Newman for the continued guidance and

support throughout the project. Additionally, I greatly appreciate the support and input

from my committee members Drs. Dan Larkin and Jacques Finlay. I would also like to

express my gratitude to the Riley Purgatory Bluff Creek Watershed District for funding

this research and providing a portion of the data used. Additionally, this thesis

incorporates data collected by John JaKa and Josh Knopik, previous UMN graduate

students who also worked in the Riley Purgatory Bluff Creek Watershed District; I am

grateful for their contributions. The Minnesota Aquatic Invasive Species Research Center

provided invaluable support for this project by supplying office and lab space. Lastly, this

work would not have been possible without the assistance from many other students,

faculty, and staff: Chloe Fouilloux, Aislyn Keyes, Todd Meilzarek, Leo Rubenstien, Matt

Gilkay, Jay Maher, Dan Krause, Dr. Sue Galatowitsch, Dr. John Fieberg, Kelsey Vitense,

Dr. Ranjan Muthukrishnan, Dr. Adam Kautza, and the staff at the University of

Minnesota Statistics Clinic.

Dedication

This thesis is dedicated to my family and friends who have continued to support my love

and curiosity for science and the natural world. My project would not have been possible

without their support.

Comprehensive Abstract

Macrophytes are a vital component to functioning aquatic ecosystems.

Specifically, macrophytes promote good water clarity by stabilizing sediments,

sequestering nutrients, and reducing the abundance of phytoplankton in the water

column. Also, macrophytes provide habitat for other aquatic organisms. Healthy, robust

aquatic macrophyte communities are indicated by diverse, abundant stands in the littoral

zones of lakes. Poor water clarity and invasive species are primary limiting factors that

cause diminished aquatic plant communities. Poor water clarity reduces the light

quantity, impeding the growth of macrophytes. Invasive fish, such as common carp

(Cyprinus carpio), damage macrophyte communities by uprooting plants and suspending

sediment and nutrients in the water column. Invasive plants, such as curlyleaf pondweed

(Potamogeton crispus) and Eurasian watermilfoil (Myriophyllum spicatum), often

outcompete native species creating dense monoculture stands.

To improve the growing conditions for macrophyte communities, several

management actions can be pursued to limit the damage by invasive species and improve

the water clarity. Common management practices in Midwestern lakes include invasive

species control and nutrient sequestration. These practices have been documented to

enhance native macrophyte communities. Lake management often requires several years

of consistent, adaptive management to effectively restore the ecosystem. Adaptive

management is the systematic process of learning from past management outcomes and

subsequently incorporating that knowledge into current management decisions. I

evaluated the change in the macrophyte community in Lake Riley, Chanhassen, MN over

the course of 6 years of lake management actions using aquatic plant point-intercept

surveys from 2011 to 2016. The results of the surveys found that after a carp removal in

2010, curlyleaf pondweed dominated the littoral zone and water clarity did not greatly

improve. Once invasive macrophytes were managed starting in 2013, incremental

increases in the species richness of the macrophyte community occurred. However,

native macrophyte expansion was limited because water clarity was still poor during the

summer growing season.

In 2016, after an alum treatment, water clarity improved and the macrophyte

community abundance and richness further increased. Species richness increased from 9

observed species in 2011 to 15 in 2016. During peak growth in August, the native species

frequency of occurrence was 50% through 2013 and then increased up to 80% of sites in

2016. The August native macrophyte biomass increased from 30g/m2 in 2011 to 600g/m2

in 2016 (p<0.05). Prior to 2016, the average maximum depth of rooted native plant

growth was 3.1m and in 2016 it increased to 4.1m. Overall, the density, coverage, and

richness of the macrophyte community increased throughout the study period

demonstrating that the macrophyte community had a positive response to the multi-year

management practices on Lake Riley.

The specific mechanism of macrophyte recruitment following improved growing

conditions, such as in Lake Riley, is an understudied area of macrophyte restoration.

Macrophytes typically propagate through clonal growth and fragmentation. However,

when macrophyte populations are reduced, the lake seed bank may contribute to the

reestablishment of the population. In previous studies on temperate lake seed banks,

seeds from vascular aquatic plants and spores from macroalgae have been found in

varying densities and viability levels suggesting that recruitment from the seed bank is

possible in some systems.

I conducted a controlled laboratory experiment using sediment from Lakes Ann

and Riley located in Chanhassen, MN, to 1) evaluate the response of the seed banks to

different treatments and 2) compare the observed taxa sprouting from the seed banks to

the taxa observed growing in the lakes. The treatments included a maximum germination

treatment using a germination promoter to evaluate the full extent of the viable seed

bank, a treatment representative of a lake with good water clarity, and a treatment

representative of a lake with poor water clarity. The good and low clarity treatments were

designed to evaluate the response of seeds to two different light levels that were observed

in lakes with high turbidity (low-light intensity) and low turbidity (high light intensity). It

was hypothesized that the maximum germination treatment would have the highest

amount of germination, the high clarity treatment would have the second highest amount,

and the low clarity treatment would have the lowest amount of germination due to the

low-light quantity.

The seed banks of both Lakes Riley and Ann were similar to the macrophyte

community observed growing in the lake. In Lake Ann, 16 species were observed

sprouting and every species observed in the experiment grew in the lake. In Lake Riley,

17 species were observed sprouting and all but two species were observed both in the

lake and in the seed bank. The seed banks did not show any significant difference in

response to the germination treatments. Chara, curlyleaf pondweed, and wild celery were

the most frequent species observed. Under maximum germination conditions, Lake Riley

had a viable vascular seed density of 2,916 ± 1,828 seeds/m2 and a viable chara spore

density of 1,033 ± 698 spores/m2. Lake Ann had a viable vascular seed density of 1,100 ±

440 seeds/m2 and viable chara spore density of 13,833 ± 2,825 spores/m2. The study

demonstrated that germinating propagules from a lake seed bank can be a valuable tool

for managers to evaluate the viable macrophyte taxa present and better understand the

potential for recruitment from the seed bank. Overall, to restore native macrophyte

communities, it requires several multi-year management actions and will likely include

multiple forms of propagule recruitment.

Table of Contents

Acknowledgements……………………………………………………………………...…i

Dedication………………………………………………………………………………....ii

Comprehensive Abstract………………………………………………………………….iii

List of Tables……………………………………………………………………………..ix

List of Figures……………………………………………………………………………..x

Prologue………………………………………………………………………………...…1

Chapter I…………………………………………………………………………………...2

Introduction………………………………………………………………………..3

The Role of Macrophytes in an Aquatic Ecosystem………………………………4

Impact of Common Carp on Macrophytes………………………………………...6

Carp Reduction and Invasive Macrophyte Response……………………………..7

Invasive Macrophyte Control……………………………………………………..9

Natural Revegetation of the Native Plant Community…………………………..10

Management Actions to Improve the Native Plant Community………………....12

Summary…………………………………………………………………………15

Literature Cited Chapter I……………………………………………………..…17

Chapter II…………………………………………………………………………...……24

Summary…………………………………………………………………………25

Introduction…………………………………………………………………...….26

Methods………………………………………………………………………......29

Study Lake……………………………………………………………….29

Invasive Species Control…………………………………………………30

Alum Treatment………………………………………………………….31

Aquatic Vegetation Surveys……………………………………………..31

Water Quality…………………………………………………………….33

Statistical Analysis……………………………………………………….34

Results…………………………………………………...……………………….36

Water Quality…………………………………………………………….36

Aquatic Vegetation Community………………………………………..37

Discussion………………………………………………………………………..42

Literature Cited Chapter II……………………………………………………….46

Tables Chapter II……………………………………………..…………………..49

Figures Chapter II………………………………………………………………..51

Chapter III………………………………………………………………………………..56

Summary……………………………………………...………………………….57

Introduction………………………………………………………………………58

Methods…………………………………………………………………………..63

Seed Bank Collection…………………………………………………….63

Seed Bank Treatments…………………………………………………...64

Sediment Analysis……………………………………………………….68

Statistical Analysis……………………………………………………….68

Results……………………………………………………………………………70

Lake Riley………………………………………………………………..70

Lake Ann…………………………………………………………………72

Discussion………………………………...……………………………………...75

Literature Cited Chapter III……………………………………………………...79

Tables Chapter III………………………………………………………………..84

Figures Chapter III……………………………………………………………….86

Chapter IV………………………………………………………………………………..90

Literature Cited Chapter IV……………………………………………………...94

Comprehensive Bibliography……………………………………………………………95

Appendix……………………………………………………………………..…………105

List of Tables

Chapter II.

Table 1. Lake Riley maximum depth of plant growth observed, percent of

native submersed taxa, the number of submersed native taxa, and the average

Secchi depth obtained from 2011 to 2016……………………………………….….…49

Table 2. Aquatic plants found in Lake Riley surveys in 2011 through 2016……….….50

Chapter III.

Table 1. Total number of propagules germinating from Lake Riley sediment for

the maximum germination, high clarity, and low clarity treatments based on

3.0L of collected sediment per treatment………………………………………….……84

Table 2. Total ungerminated propagules enumerated from the seed banks of

Lakes Ann and Riley…...……………………………………………………………….84

Table 3. Total germination of propagules from Lake Ann sediment for the

maximum germination, high clarity, and low clarity treatments based on

3.0L of collected sediment per treatment……………………………………..………...85

Appendix.

Table 1. Results from the Chapter II species richness Poisson regression

models including the R coding…………………………………………………………108

Table 2. Results from the Chapter II frequency of occurrence Poisson regression

models including the R coding…………………………………………………………109

Table 3. Results from the Chapter II biomass multiple regression models

including the R coding…………………………………………………………………112

Table 4. Results from the Chapter III Poisson regression model evaluating

the effect of treatment on total sprouts counted. Includes the R coding…………….…115

List of Figures

Chapter II.

Figure I. Riley Creek watershed map……………………………………………...…….51

Figure 2. Lake Riley Secchi depths from 2011 to 2016…………………………………52

Figure 3. Lake Riley frequency of occurrence of the most common species in

June and August surveys from 2011 to 2016…………………………………………….53

Figure 4. Lake Riley frequency of occurrence for all native species combined

and all exotic species combined from 2011 through 2016………………………………54

Figure 5. Lake Riley biomass for all native species combined and all exotic

species combined from 2011 through 2016…………………………………………...…54

Figure 6. Lake Riley biomass of the most commonly observed species in June

and August surveys from 2011 through 2016……………………………………………55

Chapter III.

Figure 1. Riley Creek watershed map……………………………………………………86

Figure 2. Lake Riley mean cumulative germination (propagules/tray) under the

different treatment conditions……………………………………………………………87

Figure 2. Lake Riley average germination (seeds/tray and spores/tray) under the

maximum germination, high clarity and low clarity treatment conditions for the

most abundant species observed…………………………………………………………87

Figure 4. Lake Ann mean cumulative germination (propagules/tray) under the

different treatment conditions……………………………………………………………88

Figure 5. Lake Ann average germination (seeds/tray and spores/tray) under the

maximum germination, high clarity and low clarity treatment conditions for

the most abundant species observed……………………………………………………..88

Figure 6. Lake Ann seed bank and point-intercept survey rankings scatterplot…………89

Figure 7. Lake Riley seed bank and point-intercept survey rankings scatterplot………..89

Appendix.

Figure 1. Lake Riley Secchi depths from 2010 to 2016………………………………..106

Figure 2. Lake Riley temperature and dissolved oxygen from 2010 to 2016………..…107

1

Prologue

There are four chapters in this thesis. Chapter I is a summary of the role of

macrophytes in a lake ecosystem and common lake management actions to improve

aquatic macrophyte communities. Chapter II assesses the response of the macrophyte

community in Lake Riley, Chanhassen, MN, U.S.A. to several lake management actions

including common carp (Cyprinus carpio) reduction, invasive macrophyte control, and an

alum treatment. Using point-intercept surveys, I evaluated the change in species richness,

abundance, and biomass over time using results from 2011 to 2016. Once data from the

summer of 2017 are incorporated, the chapter will be submitted to a journal for

publication.

In Chapter III, I evaluated the effect of water clarity and subsequently light

quantity on the germination of propagules (seeds/spores) from lake sediments from Lakes

Riley and Ann, Chanhassen, MN. I exposed seed bank samples to one of three

treatments: maximum germination, high clarity, and low clarity. I also compared the

viable seed bank to what was observed growing in the lakes during point-intercept

surveys. This chapter will also be submitted to a journal for publication.

Chapter IV serves to summarize the entire thesis and link the main concepts

between the chapters and compare these results to other similar studies. Overall, this

thesis provides insight on the macrophyte response to lake management actions and the

potential role of the lake seed bank in the revegetation of a lake after growing conditions

are improved.

2

Chapter I

The Restoration of Aquatic Macrophyte Communities in Temperate Lakes

3

Introduction

Shallow lake restoration projects often begin with turbid, phytoplankton

dominated lakes with the aim to flip the ecosystem to a clear-water, macrophyte

dominated lake (Hupfer et al. 2016, Stroom and Kardinaal 2016). Aquatic macrophyte

restoration is a multi-step process requiring adaptive management and long-term

planning. Adaptive management is the systematic process of learning from past

management outcomes and subsequently incorporating that knowledge into current

management decisions (Westgate et al. 2013).

Currently, there are several methods used to improve the lake conditions for

macrophyte regeneration. Improved water clarity is often pursued first as light is a

primary limiting factor for plant growth (Bornette and Puijalon 2011, Verhofstad et al.

2016). Water clarity improvement is achieved by reducing nutrient concentrations within

the lake because high nutrient concentrations sustain the growth of phytoplankton.

Actions to minimize nutrients include decreasing the external loading of nutrients into the

system and biomanipulation, such as reducing the benthivorous fish population or

altering the trophic structure of the food web (Cooke et al. 2016). A reduction of internal

phosphorus loading may also be needed if the total phosphorus within the system is still

high (Hupfer et al. 2016).

Improved water clarity can aid in macrophyte expansion and may enhance

propagule (seeds and spores) germination conditions for several species. However, it can

also allow for exotic, low-light tolerant species to establish in high density before many

native species. Therefore, as water clarity improvements are pursued, the active control

4

of invasive species such as curlyleaf pondweed (Potamogeton crispus) and Eurasian

watermilfoil (Myriophyllum spicatum) may also be necessary to allow for the expansion

of the native macrophyte community.

The Role of Macrophytes in an Aquatic Ecosystem

Macrophytes include submersed and emergent aquatic vascular plants in addition

to macroalgae such as those from the order Charales (Cooke et al. 2016). Macrophytes

play an important role in maintaining the water clarity of a lake ecosystem. Aquatic

macrophytes use nutrients such as nitrate and phosphate, limiting the amount of nutrients

available for phytoplankton uptake (Dennison et al. 1993). Additionally, they can retain

the sediment, which further sequesters nutrients and reduces turbidity (Dennison et al.

1993, Horppila and Nurminen 2003). As macrophytes effectively compete with

phytoplankton for nutrients in the water column, algal blooms and the associated

turbidity are limited, preventing decreases in water quality (Bakker et al. 2010).

A robust macrophyte community is an integral part of a functioning aquatic

ecosystem as it also creates habitat for aquatic organisms. Zooplankton can thrive within

stands of vegetation that act as a shelter and protects them from predation (Donk and

Bund 2002). Zooplankton, such as Daphnia spp., in turn also contribute to good water

clarity by consuming phytoplankton, including blue-green algae species (Schoenberg and

Carlson 1984). Juvenile fish rely on the complex structure provided by diverse stands of

macrophytes as protection from piscivorous fish and as a food resource, consuming

zooplankton and macroinvertebrates located in the plant beds (Valley et al. 2004, Cross

5

and McInerny 2006). Moreover, mature piscivore populations have been shown to be

effective at controlling planktivorous fish populations, which are known to consume

zooplankters (Drenner and Hambright 2002). As a result, the piscivorous fish serve as a

top down control and assist in maintaining high zooplankton populations, thus sustaining

good water clarity (Drenner and Hambright 2002). Overall, a diverse macrophyte

community increases the complexity of habitat and subsequently increases the diversity

of present aquatic life.

These trophic relationships can result in a lake settling into one of two alternative

stable states: a clear water, macrophyte dominated lake and a turbid water, phytoplankton

dominated state (Scheffer et al. 1993). These ecosystem states are considered to be stable

because lake systems tend to resist change due to reinforcing feedbacks within the

ecosystem. Change to the other state occurs when perturbation to the ecosystem reaches a

critical threshold disrupting the reinforcing feedbacks (Scheffer and Ness 2007).

Common perturbations in lakes include large benthivorous fish populations and

eutrophication (Scheffer et al. 2001).

Several studies have demonstrated that lake management can be successful in

maintaining the clear-water stable state as opposed to the phytoplankton dominated,

turbid stable state (Scheffer et al. 1993, Hansson et al. 1998). Macrophyte expansion is

often necessary to maintain the clear-water alternative stable state due to their role in

sequestering nutrients, suppressing sediment suspension and providing habitat for

herbivorous zooplankters (Hansson et al. 1998, Hilt et al. 2006). To attain a clear-water

state, a reduction in external loading of nutrients is essential. Excess nutrient loading into

6

the system can fuel phytoplankton growth, which limits the water clarity and ability for

macrophytes to grow within the littoral zone (Schindler 2006). To effectively shift the

lake to a clear-water state following a reduction in external loading, several actions may

be pursued to establish water clarity conditions that will allow for macrophyte growth

and expansion in the littoral zone (Hilt et al. 2006). Potential actions include

benthivorous fish reduction, decreasing internal nutrient loading, and invasive

macrophyte control.

Impact of Common Carp on Macrophytes

The invasion of common carp (Cyprinus carpio) throughout the U.S. has greatly

diminished water clarity and macrophyte assemblages in lakes across the country (Weber

and Brown 2009). Common carp are a benthivorous fish that can have a myriad of

cascading effects on lake ecosystems (Kloskowski 2011). Documented cases in

Midwestern lakes have demonstrated that when carp reach a threshold biomass of

approximately 100kg/ha, they can cause the switch to a phytoplankton dominated, turbid

stable state (Bajer et al. 2009). Carp promote this condition through their feeding

behavior. They consume benthic plants and sediment and as such cause the suspension of

sediment and nutrients into the water column (Crivelli 1983). In addition, this behavior

causes the uprooting of aquatic plants (Crivelli 1983). With the increase in nutrients and

the decrease in macrophytes, algal blooms become common and phytoplankton can

dominate the system (Zambrano et al. 2001). Some additional research has also

demonstrated that carp may have a profound impact on the lake seed bank. Carp may

7

consume seeds and spores, with complex outcomes for seed banks, inducing germination

in some species while decreasing the viability of propagules in others (Pollux 2011).

Overall, carp decrease water clarity and significantly reduce the ability of macrophytes to

establish.

Carp Reduction and Invasive Macrophyte Response

In the Midwest region of the U.S., carp removals have been completed to reduce

the biomass of the population and to begin the shift to the macrophyte dominated, clear-

water stable state. Effective benthic fish removals have been documented to reduce

sediment and nutrient suspension in the water column (Hanson and Butler 1994, Bajer

and Sorensen 2015). Following reduction in carp biomass below the critical threshold,

macrophytes have been documented to re-establish from intact source populations (De

Backer et al. 2012). Also, the natural recruitment of vegetation from the seed bank has

been documented to occur if the seed bank is still viable (De Backer et al. 2012, Knopik

2014). The extent of macrophyte recolonization following clarity improvement is

variable among lakes, and macrophyte coverage in the years following fish removal

ranged from 20% to 90% of the lake area (Reynolds 1994, Norlin et al. 2005, Knopik

2014).

Following reduction in external loading and fish removal, there is a risk of

invasive vegetation proliferating under the new conditions before natives can expand or

sprout from the seed bank. Low-light tolerant species can often dominate the system as

water clarity slowly improves before most species are capable of germination and growth

8

(Lauridsen et al. 1993, Jeppsen et al. 2005, Søndergaard et al. 2008). Invasive

macrophytes in North America, such as curlyleaf pondweed (Potamogeton crispus) and

Eurasian watermilfoil (Myriophyllum spicatum), are adapted to grow under low-light

conditions and in cold temperatures enabling them to establish early in the growing

season and to grow in deeper locations (Nichols and Shaw 1986).

Curlyleaf pondweed often sprouts from vegetative buds called turions, which are

deposited on the lake bed in mid-summer and store energy for the plant to grow under the

ice in winter and early spring (Nichols and Shaw 1986). By the early summer, curlyleaf

can form dense, monospecific stands that impede recreation and greatly alter the

ecosystem (Bolduan et al. 1994). Due to this early season growth, many native species

are unable to sprout from propagules or germinate from the seed bank at the appropriate

time due to the shading imposed by the curlyleaf beds (Santos et al. 2011). Additionally,

curlyleaf senesces in mid-summer providing a source of phosphorus for phytoplankton

that can form algal blooms (Bolduan et al. 1994).

Eurasian watermilfoil can also form dense stands, matting out at the surface and

shading other vegetation that may be growing below (Smith and Barko 1990, Madsen et

al. 1991). Eurasian watermilfoil propagates largely through fragmentation and seeds are

thought to play a minimal role in the spread of the species (Madsen et al. 1988). A low-

light and cold tolerant species, Eurasian watermilfoil can also be observed growing under

ice, shading out natives later in the growing season (Nichols and Shaw 1986).

9

Invasive macrophyte control

After the shift to the clear-water stable state has begun, through biomanipulation

and external loading reduction, control of invasive macrophytes may be needed. Invasive

aquatic macrophytes are easily spread and can occur in damaging abundances (Pimentel

et al. 2005). There are several treatment methods that can be employed to control

invasive or nuisance levels of macrophytes. Herbicide applications have been

documented to effectively control, not eradicate, populations of curlyleaf pondweed and

Eurasian watermilfoil. Specifically, early season endothall treatments have been

demonstrated to target curlyleaf pondweed while it is actively growing and not

significantly impair the growth of native macrophytes, mainly because they have not yet

started to grow (Johnson et al. 2012, Jones et al. 2012, JaKa 2015). Similarly, Eurasian

watermilfoil has been controlled by the use of 2,4-D and triclopyr herbicides, which

primarily affect dicot species and not the majority of native macrophytes, which are

monocots (Cooke et al. 2016). Therefore, selective control measures that target the

prolific invasive species are highly useful in also maintaining an intact native plant

community. The level of impact on native macrophytes is dependent on the dose

concentration as well as the length of exposure. The use of herbicides is highly regulated

and testing has determined limited impacts on non-targeted aquatic life (Cooke et al.

2016). Herbicide applications often require multiple years of treatment and the longevity

of treatments is still an area of ongoing research (Nault et al. 2014, JaKa 2015,

Netherland and Jones 2015).

10

Other methods used to control invasive macrophyte species include harvesting,

dredging, and sheeting (Cooke et al. 2016). However, these methods tend to not be as

species specific as the selective herbicide applications and they can affect the native plant

community as well. Certain methods may work better than others depending on the lake

system, therefore, the need for adaptive management and long-term planning are needed

when managing invasive macrophytes.

Natural Revegetation of the Native Plant Community

Some cases have demonstrated that after biomanipulation and selective control of

invasive macrophytes, native plants can increase in frequency and biomass due to the

increases in clarity and reduced competition (Lauridsen et al. 1993, Getsinger et al. 2000,

Jones et al. 2012, JaKa 2015). To establish a diverse aquatic plant community

environmental conditions such as light, temperature, nutrients, and sediment

characteristics must meet germination and growth demands (Arthaud et al. 2012). There

are several ways for macrophyte communities to naturally reestablish before the need for

further manipulation occurs.

Submersed aquatic vegetation often colonizes from a source population;

fragmentation or clonal growth are the most common form of spread for aquatic

vegetation (Boedeltje et al. 2002, Boedeltje et al. 2003). Sources of fragments may be

from a small population that remained in the lake, from upstream drift, or via human or

animal transport between lakes (De Winton et al. 2000, Figeroula and Green 2002). The

role of the seed bank is relatively minimal compared to the spread of species by

11

fragmentation or clonal growth (Santamaría 2002). However, when macrophyte species

are eliminated from a lake system, through benthivorous fish disruption or competition

with invasive plants, the seed bank may play an important role in recolonization (De

Winton et al. 2000). Macrophyte propagules (seeds and spores) can also be spread

between lake systems by wind-induced currents in a lake or by animal transport (Brochet

et al. 2010).

There have been several studies assessing the role of the seed bank in revegetating

wetland ecosystems, however few studies have been completed to assess the role of the

seed bank in lake ecosystems (Haag 1983, McFarland and Schafer 2011). The studies that

have occurred demonstrate that the present vegetation in a wetland or lake does not tend

to correlate with the assemblage of seeds in the seed bank (Titus and Hoover 1991,

Combroux and Bornette 2004). Lake sediments are also known to have lower propagule

densities than wetland and riparian areas and the length of propagule viability is also

highly variable among species and lake environments (Kleyer et al. 2008, De Backer et

al. 2012). The timing of propagule germination in lake sediments has not been

thoroughly investigated although it is thought to occur in late spring and early summer as

water temperature and day length increases, likely triggers to break dormancy (Baskin

and Baskin 2014).

The germination of propagules from the seed bank in a lake requires several

environmental factors, including light (Sederias and Colman 2007). Light induced

germination is common for macrophyte propagules and studies have demonstrated that

burial reduces the germination, potentially due to lack of light (Dugdale et al. 2001,

12

Baskin and Baskin 2014). In some systems, sediment disturbances are needed to suspend

propagules to the top of the lake sediment to optimally induce germination in clear water

systems where light can trigger germination cues. Other factors important to germination

in some macrophyte taxa include temperature and dissolved oxygen (Dugdale et al. 2001,

Arthaud et al. 2012).

Management Actions to Improve the Native Plant Community

When the return of native macrophyte species does not occur after a

biomanipulation and reduction in external loading, low water clarity is often determined

to be the cause impairing the ability of plants to grow and establish a robust population

(van de Hatered et al. 2007). There are several reasons why water clarity may not

improve enough for the establishment of a healthy macrophyte community. Internal

phosphorus loading may still cause excessive phytoplankton blooms, maintaining a turbid

state during the growing season. Alternatively, zooplankton may not recover quickly to

high densities thus failing to control the phytoplankton growth. Additionally, the seed

bank may be exhausted if carp have consumed a significant portion of propagules from

the seed bank or there has been a lack of replenishment due to many years of suppressed

macrophyte communities (Pollux et al. 2006, Vojtko et al. 2017). Even if clarity

improves relative to the pre-carp removal levels, it may not be a sufficient amount to

allow for substantial macrophyte growth or to induce germination in the seed bank and

thus still be a limitation on the expansion of macrophytes throughout the lake and in

deeper areas of the littoral zone (Chambers and Kalff 1985). When light conditions are

13

not sufficient to allow for the recolonization of macrophytes after invasive species

control, methods to improve the water clarity can be undertaken to improve the growing

conditions.

A common tool used by lake managers to increase the water clarity is to reduce

the internal loading of phosphorus in the lake. Internal loading of phosphorus is a

common issue in lakes that stratify and become anoxic in the sediment for part of the

year. Phosphorus is bound to ferric iron within the lake under oxygenated conditions

(Cooke et al. 2016). When a stratified lake becomes anoxic, redox reactions in the

sediment reduce iron, from the ferric to ferrous state, which no longer binds phosphorus.

The phosphorus thus becomes mobilized again in the sediment and water column,

available for uptake by organisms (Cooke et al. 2016). Thus, alternative approaches to

bind phosphorus are used in eutrophic lakes such as an alum treatment.

Alum is an aluminum salt, which when added to water creates a floc that binds to

phosphorus in the water column as it settles to the lake bottom and also binds to

phosphorus in the sediment when it becomes mobilized again due to iron reduction

(Barko et al. 1990). By conducting an alum treatment, the amount of available

phosphorus due to internal loading is decreased and water clarity is improved due to a

reduction in phytoplankton. Successful alum treatments have demonstrated an increase in

water clarity and a decreased internal loading rate, total phosphorus concentration, and

chlorophyll-a concentration within the system (Barko et al. 1990, Welch and Sherieve

1994, Welch and Cooke 1999, Huser et al. 2011).

14

The increase in water clarity after an alum treatment has been associated with an

increase in macrophytes due to greater light availability. However, macrophytes can grow

to nuisance levels and the increase is often due to exotic species (Welch and Shrive 1994,

James 1996). One study in Minnesota demonstrated a successful alum treatment with an

improvement in clarity however the vegetation increases were due to Eurasian

watermilfoil (Huser et al. 2011). Currently, there are few studies that assess the capacity

of the seed bank to respond to the increase in water clarity provided by an alum

treatment. If seed bank sampling occurs and there appears to be viable propagules of

desirable species, then one can monitor the effect of increased clarity and assess the

extent of the propagule response. If sampling indicates few viable species, potentially due

to benthivorous fish populations or poor sediment conditions, alternate measures will

likely to be needed to aid in the recovery of the population if there is no source of

recruitment for new species within the system.

The return of diverse macrophyte stands may be impeded by an impoverished

regional pool of propagules, in this case there are additional restoration practices that can

be implemented (Sand-Jensen 2008, Dudley et al. 2012). In this case transplanting of

macrophytes from nearby systems may be a viable option (Smart et al. 1998). Following

reintroduction, some species may establish and spread, called the founder colony

approach (Smart et al. 1998, Cooke et al. 2016). However, several species do not survive

transplanting experiments (Smart et al. 1998). The biggest factor attributed to limited

success is poor water clarity (Knopik 2014). With limited light conditions macrophytes

cannot expand. Therefore, improving water clarity still appears to be a vital step in the

15

promulgation of a native plant community prior to beginning transplanting. One method

of inducing native plant growth in a newly improved system that has not been

investigated is transferring sediment and the seed bank from species rich systems to

systems with a diminished seed bank. The propagules may be induced to germinate under

the clear water conditions if spread over the current sediment, however there are concerns

about unintentional species introductions. This is an area of research that requires further

investigation.

Summary

Several environmental factors are important in the establishment or restoration of

a native macrophyte community in lake ecosystems. In some cases, after reduction in

external loading and biomanipulation, the clear water state is not improved to the degree

that native macrophytes can repopulate the lake, leaving low-light tolerant species such

as curlyleaf pondweed and Eurasian watermilfoil able to establish and become

widespread at nuisance levels. Management of invasive macrophytes is common and

often effective, however native plant communities still may not rebound. To aid in the

restoration of a macrophyte community, further improvements in water clarity are often

pursued next by the reduction of internal loading. Natural revegetation of some native

plant species has been documented to occur from fragmentation and the seed bank. If the

seed bank and available propagules are low, transplanting of species is a viable next step.

Overall, the establishment of a stable, diverse submersed aquatic plant community

requires several management actions to reduce nutrient loading, sediment suspension, and

16

to control invasive macrophytes before revegetation efforts are made if they are not

naturally occurring as the clarity improves. Once an established macrophyte population

exists, the clear water state may become stabilized.

17

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24

Chapter II

The Restoration of Native Aquatic Macrophytes: Macrophyte Response to Carp

Reduction, Invasive Macrophyte Control, and Alum Treatment

25

Summary

I evaluated the response of the aquatic vegetation community to several

management actions in Lake Riley, Carver County, MN (MN DNR DOW ID 10-000200)

using surveys from 2011 through 2016. Management actions included common carp

(Cyprinus carpio) removal, selective control of invasive macrophytes, and an alum

treatment. Using the point-intercept survey method, the frequency of occurrence and

biomass of plants was assessed in June and August each year and the changes were

evaluated using regression models. The changes in the native and invasive macrophyte

species richness and abundance were also evaluated using regression models.

Following carp removal in 2010, the plant community was largely dominated by

curlyleaf pondweed and water clarity did not greatly improve relative to pre-carp removal

levels. Species richness increased gradually over the surveyed years as invasive species

were managed with herbicide control starting in 2013. Water clarity significantly

improved due to the alum treatment in May 2016 and the macrophyte community

richness and abundance increased further. Species richness increased from 7 in 2011 to

15 in 2016. Additionally, throughout the study, the native species frequency of

occurrence and biomass increased. During peak growth in August, native species

frequency of occurrence was 50% in 2011 to 2013 and then steadily increased up to 80%

of sites in 2016. The August native macrophyte biomass increased from 30g/m2 in 2011

to 600g/m2 in 2016 (p<0.05).

Following the alum treatment, the native plant community grew in deeper waters

relative to pre-alum treatment observations. The maximum depth of rooted native plant

26

growth observed in 2016 was 4.1m, whereas prior to 2016 the average maximum depth of

rooted native plant growth was 3.1m. Lastly, invasive macrophytes, curlyleaf pondweed

and Eurasian watermilfoil, were controlled by the herbicide treatments and the frequency

of occurrence and biomass of these taxa decreased relative to pre-control levels and

remained low despite water clarity improvements. Overall, the management actions in

Lake Riley resulted in a denser, more diverse macrophyte community. This case study

demonstrates the need for adaptive, multi-year management when restoring macrophyte

communities and lake ecosystems.

Introduction

Light availability is an environmental factor associated with the germination and

survival of submersed aquatic macrophytes. When water clarity is poor many macrophyte

species do not obtain enough light to efficiently photosynthesize (Binzer et al. 2006,

Rodrigues and Thomaz. 2010). Often in low clarity lakes the macrophyte community has

low species diversity and is dominated by low-light tolerant species (Binzer et al. 2006,

Cooke et al. 2016). In central Minnesota lakes, poor water clarity is typically due to

eutrophication and large populations of benthivorous fish. Eutrophication is caused by

high nutrient concentrations resulting in the proliferation of phytoplankton in the water

column (Schindler 2006). High phytoplankton abundance increases light attenuation and

subsequently reduces the capacity for macrophytes photosynthesize (Binzer et al. 2006,

Cooke et al. 2016). Additionally, common carp are widespread in Minnesota lakes,

impairing water quality and macrophyte communities. Common carp suspend sediment

and nutrients by spawning in macrophyte beds and uprooting macrophytes as they feed

27

on macroinvertebrates in the sediment. (Bajer et al. 2009). These behaviors increase the

turbidity and nutrient concentrations in the water column, subsequently also increasing

the abundance of phytoplankton (Bajer et al. 2009).

Diminished macrophyte populations have significant impacts on other aquatic

organisms as macrophytes play an integral role in the functioning of a lake ecosystem.

Macrophytes sequester nutrients and sediment promoting water clarity (Dennison et al.

1993). Additionally, macrophytes provide rearing habitat for invertebrates and fish (Donk

and Bund 2002, Cross and McInerny 2006). The lack of macrophytes in some Minnesota

lakes have resulted in turbid water, higher nutrient concentrations, and a change in the

composition of aquatic organisms as exotic species proliferate throughout the littoral

zone (Hilt et al. 2006, Knopik 2014, JaKa 2015).

If the macrophyte population in a lake system is in low abundance and not

diverse, lake managers can pursue several strategies to improve the growing conditions

(Van de Hatered et al. 2007). Firstly, reducing the external loading of nutrients and

sediment into the aquatic ecosystem is a critical management practice. This can be

achieved by implementing riparian and upstream projects that aim to limit the amount of

sediment and nutrient inputs into the system, such as reducing fertilizer use or planting

riparian vegetation (Hilt et al. 2006). Secondly, if the benthivorous fish population is in

high abundance, reduction of the population may be necessary to limit the damage to

growing macrophytes and prevent the increased release of sediment and nutrients into the

water column (Bajer et al. 2009).

28

If a reduction in external loading and benthivorous fish control do not yield an

improvement in water clarity, a reduction in internal loading may be necessary to

decrease the nutrient concentrations and phytoplankton abundance. This can be achieved

by applying aluminum salts (alum) to the water, which bind to mobile phosphorus in the

water column and sediment. Documented cases demonstrate that macrophyte

colonization can increase following an alum treatment however, invasive macrophytes

may respond more rapidly than native macrophytes (Newman et al. 2004, Spears et al.

2016). Therefore, management of invasive aquatic plants is often necessary as water

clarity improves throughout the restoration process (Newman et al. 2004, JaKa 2015).

Common management techniques for invasive macrophyte control include the use of

mechanical removal such as a harvesting or the use of selective herbicides approved for

aquatic use.

Several studies have evaluated the response of macrophyte communities to lake

management actions. Some examples include the evaluation of the effects of endothall

herbicide on macrophyte communities (Jones et al. 2012, JaKa 2015), the effect of

reducing internal nutrient loading on macrophyte abundance (Newman et al. 2004,

Spears et al. 2016), and the effect of carp reduction on the macrophyte community (Bajer

et al. 2009). However, fewer studies evaluate the response of the macrophyte community

to multiple management actions occurring over several years. Most studies have been

three years or less in duration and do not evaluate long-term management strategies (Hilt

et al. 2006, Johnson 2011, JaKa 2015, Cooke et al. 2016).

29

The purpose of this study was to assess the native and non-native submersed

macrophyte community after carp removal from 2011 to 2016 in response to management

actions. I assessed if selective invasive macrophyte control and an alum treatment can be

used to increase the abundance and richness of a macrophyte community in a lake after

successful carp reduction. After carp removal in 2010-2011, Lake Riley underwent

curlyleaf pondweed control in early spring each year from 2013 through 2016 with

endothall herbicide applied in delineated treatment blocks. Eurasian watermilfoil was

treated in June of 2015 and 2016 using 2,4-D herbicide in delineated treatment blocks. A

hypolimnetic alum treatment was conducted in May of 2016. Overall, this study assessed

the change in abundance and diversity of the plant community in Lake Riley using point-

intercept surveys that occurred each June and August from 2011 through 2016. Long-

term, adaptive management is necessary for macrophyte restoration, and yet it is rarely

assessed. This study provides insight into the key management factors that influence

macrophyte communities.

Methods

Study Lake

Lake Riley (DOW 10-0002) is a eutrophic lake located within the Riley Purgatory

Bluff Creek Watershed District (RPBCWD) in Chanhassen and Eden Prairie, Minnesota

USA. Lake Riley has a maximum depth of 15m and is a dimictic lake, stratifying during

the winter and summer. Lake Riley is 120 hectares in area and is within the Riley Creek

drainage, with a watershed area of 2,590 hectares (Figure 1). The land use in the Riley

30

Creek watershed is suburban with a mixture of residential housing, commercial

infrastructure, farmland and forested parks.

There were several management actions pursued on Lake Riley over the last 7

years. Common carp were removed from Lake Riley in 2010, as part of an attempt to

improve water clarity and quality (Bajer et al. 2011). Following carp removal, water

clarity marginally increased and limited native macrophyte expansion occurred (JaKa

2015). The invasive species curlyleaf pondweed and Eurasian watermilfoil were

controlled in Lake Riley by multi-year, early season herbicide applications of endothall

(2013-2016) and 2,4-D (2015-2016) respectively. In 2016, a hypolimnetic alum

treatment was conducted to reduce internal nutrient loading and improve water clarity.

Invasive Species Control

To reduce the common carp population in Lake Riley, carp were removed in 2010

by winter seining. Prior to the removal, the estimated biomass of carp was 176.1 kg/ha in

2009. After the reduction occurred, the estimated biomass was reduced to 90.0 kg/ha

(Bajer and Sorensen 2012). Throughout the study period, the carp population remained

below 100 kg/ha in Lake Riley because of lack of recruitment and active control of the

population after the initial removal.

To control curlyleaf pondweed, areas of dense growth were delineated in Lake

Riley in early spring in 2013, 2014, 2015, and 2016. These areas were determined both

visually and by throwing a double-headed garden rake over the boat to confirm presence

or absence of curlyleaf pondweed. Treatment blocks were delineated using ArcGIS and

31

when the water temperature reached 10 to 15 ˚C the delineated blocks were treated with

Aquathol K ®, a 40.3% dipotassium salt of endothall herbicide (7-

oxabicyclo[2,2,1]heptane-2,3-dicarboxylic acid) at a targeted concentration of 1.0mg/L.

The herbicide application is selective for curlyleaf pondweed because most native plants

are still dormant at this temperature and thus not affected (Johnson et al. 2012, Jones et

al. 2012). Herbicide was applied to approximately 8 hectares in 2013, 13 hectares in

2014, 8 hectares in 2015 and 7hectares in 2016.

For control of Eurasian watermilfoil, the delineation of dense areas of growth

occurred in late May in both 2015 and 2016 using the same methods as the curlyleaf

pondweed delineation. Delineated treatment blocks were treated in mid-June in 2015 (14

hectares) and 2016 (13 hectares) with 2, 4- Dichlorophenoxyacetic acid (2, 4-D)

herbicide at a targeted concentration of 2.0mg/L. The herbicide is selective for dicot plant

species, therefore it is primarily selective for Eurasian watermilfoil.

Alum Treatment

Lake Riley received a hypolimnetic alum treatment (approximately 100 hectares)

in early May 2016. The application occurred using a customized boat that followed GPS

coordinates and released the alum (aluminum sulfate) into the water column. When

applied, the alum creates a floc that binds to mobile phosphorus in the water column and

in the sediment (Cooke et al. 2016). By sequestering the available phosphorus, the alum

floc reduces the availability to phytoplankton thus decreasing the turbidity in the water

column.

32

Aquatic Vegetation Surveys

Point-intercept surveys (Madsen 1999) were conducted to measure the frequency

of occurrence of macrophyte species in Lake Riley in June and August between 2011 and

2016. Points were randomly created on a 50m grid in the littoral zone using ArcGIS.

There were 185 survey points in the littoral zone sampled during each survey. The littoral

zone is legally defined by the Minnesota Department of Natural Resources as depths less

than or equal to 4.6m. The points were uploaded onto a Garmin GPSmap76 GPS. Once

on the lake, a boat was navigated to each survey point using the GPS.

To obtain the macrophyte frequency of occurrence, at each survey point a double-

headed metal garden rake connected to a rope was tossed and then allowed to sink to the

lake bed. The rake was dragged along the lake bottom approximately 10 meters and then

pulled to the surface when it reached the boat. The depth, species present, relative

abundance, and overall rake density was observed. The rake density rating was on a scale

of zero to five based on the extent that plants filled the rake, with five equating to a full,

dense rake and one equating to a sparse rake. Empty rakes were given a rating of zero.

The plants on the rake were identified to species and each species received an abundance

rating from zero to five.

Plant biomass samples were also collected during point-intercept surveys. A

subsample of 40 points were randomly selected for biomass sample collection. A single

headed garden rake was used to sample an area of 0.09m2 (Johnson and Newman 2011).

The 0.33m garden rake was lowered to the lake bottom and rotated three times and

retrieved to obtain all the plants in the sampling area (Johnson and Newman 2011).

33

Biomass samples were placed in sealable plastic bags and stored in a cooler for transport

to the laboratory. At the laboratory, samples were stored at 5 °C until they were

processed and sorted. Samples were rinsed to remove sediment and excess debris and

remaining roots were removed to obtain an estimate of above ground biomass. The

separate species were spun in a salad spinner to remove excess water prior to drying and

then weighed to obtain a wet biomass measurement. The separate species were then

placed in pre-weighed brown paper bags. Plants were dried for at least 48 hours at 105°C

and reweighed. Plant biomass was calculated as grams dry per square meter (g dry/m2) by

dividing the dry sample mass by the total sample area (0.09m2). Mean lake-wide littoral

plant biomass was calculated by averaging all samples from depths of ≤ 4.6m for each

individual plant species and for total native biomass including points where plants were

not present.

The results of the point-intercept surveys (frequency of occurrence) and biomass

sampling are reported as an average from the June and August surveys.

Water Quality

A set of water quality measurements were recorded in Lake Riley at the deepest

part of the lake at midday throughout the growing season from May through August and

during the point-intercept surveys. Dissolved oxygen (DO), temperature, and

photosynthetically active radiation (PAR) were measured at 0.5m increments until the

hypolimnion was reached or until the cable ran out at 9.0m. DO and temperature were

measured using a YSI ProODO electronic meter and measured in mg/L and °C

34

respectively. Values for PAR were measured using a LiCor Li-189 Light Meter and a Li-

Cor underwater quantum sensor and recorded in µmol photons/s/m2. Secchi depths

(nearest 0.1 m) were also recorded during water quality assessments and compared

between years. Additionally, staff from the Riley Purgatory Bluff Creek Watershed

District took water quality measurements every two weeks from April to October on

Lake Riley during the study. The total phosphorus measurements obtained by the staff

members from 2013 to 2016, were compared between years to evaluate the effectiveness

of the alum treatment.

Statistical Analysis

All statistical analysis was conducted using R statistical software version 3.3.2

(The R Foundation for Statistical Computing, 2016). Results were considered statistically

significant when p values were < 0.05. The frequency of occurrence data followed a

Poisson distribution. Resultantly, Poisson regression models were used to evaluate

significant factors affecting macrophyte frequency of occurrence for both June and

August surveys. Factors included in the model to evaluate total native species abundance

and individual native species abundance were year, month, exotic species frequency of

occurrence, and pre- or post-alum treatment. Year was used to capture the climatic

variability between survey years. Month was used to account for the difference in

abundance in the early season (June) compared to the late season (August). To evaluate

the change in the exotic species frequency of occurrence in both June and August

(Eurasian watermilfoil and curlyleaf pondweed), Poisson regression models were also

35

used and included the following parameters: year, month, native macrophyte frequency

of occurrence, and pre- or post-herbicide treatment.

Multiple regression models were used to detect significant factors influencing

total mean native species biomass and the mean biomass of each individual species in

both the June and August surveys. Factors evaluated included year sampled, month, pre-

or post-alum treatment, and mean exotic species biomass. Mean exotic species biomass

(Eurasian watermilfoil and curlyleaf pondweed) were also evaluated for June and August

surveys and the models included year, month, pre- or post- aquatic plant herbicide

control, and total mean native species biomass. The log plus 1 of the biomass values was

used to account for the non-normal distribution of the mean biomass values and also

biomass values of zero.

Initially, mixed effects models were tested to account for the sampling point as

the random effect and the alum treatment as the fixed effect. However, the mixed effect

models resulted in a conservative estimate of variability and do not account for the fact

that when we resampled points each year during point-intercept surveys, we did not

precisely return to the same point due to GPS accuracy and human sampling error.

Therefore, implementing multiple regression models allowed for a more precise

assessment of the change in the plant community and accounts for other environmental

factors such as year and the presence and density of other species.

To assess the significance in the change in August species richness, a Poisson

regression model was used to assess the effect of the alum treatment, exotic species

present, month, and year on the total August richness observed in the lake. Additionally,

36

the change in the species richness per point in August was evaluated based on the number

of exotic species at the site, alum treatment, month, and year. Only August richness was

evaluated because that is the peak growth for native macrophyte species (Cooke et al.

2016). Lastly, a linear model was used to assess the change in the depth of rooted native

macrophyte growth using alum treatment, exotic species biomass, month, and year to

evaluate the change in colonization depth.

Overall, these statistical models were implemented to assess the change in the

macrophyte community over the survey period due to the identified management factors

of alum treatment and invasive macrophyte control. Carp biomass was not included as a

factor because no pre-treatment data exists on the macrophyte community.

Results

Water quality

In May, the average Secchi depth was variable throughout all survey years. Prior

to the 2016 alum treatment, the May Secchi depth varied between 1.1m and 3.5m from

2011 to 2015. After the alum treatment occurred on May 9th and 10th of 2016, the Secchi

depth was 6.5m (Figure 2, Appendix Figure 1). Before the alum treatment, Secchi depths

decreased to an average of 1.2m in July and 0.5m in August. After the alum treatment,

the average August Secchi depth was 1.8m. During all years of monitoring, from 2011

through 2016, the dissolved oxygen and temperature profiles generally showed an anoxic

hypolimnion below 5.0m in August (Appendix Figure 2). The August thermocline was

consistently between 5.0m and 6.0m during all survey years (Appendix Figure 2). The

37

PAR was variable throughout the survey years due to different levels of cloud cover on

survey days. The total phosphorus (TP) levels decreased following the alum treatment.

August TP levels averaged at 0.055mg/L ± 0.012mg/L prior to the alum treatment. After

the alum treatment, the TP levels decreased in August to 0.048mg/L ± 0.009mg/L.

Aquatic vegetation community

The native macrophyte community steadily increased in richness and abundance

after the carp removal, exotic species treatments, and the alum treatment on Lake Riley.

In 2011, immediately after carp removal, the August native species richness in Lake

Riley was 7 species. Following the curlyleaf pondweed and Eurasian watermilfoil

herbicide treatments the August native richness increased to 11 species in 2015, and after

the alum treatment in 2016, 14 native species were observed in August (Table 1). The

increase in lakewide August species richness was not significant (Appendix Table 1).

However, the mean species richness per sampling point increased from 0.6 species in

August of 2011 to 1.6 species in August of 2016 and this increase was significant

(p<0.01). The results of the Poisson regression demonstrated that the August species

richness per point was significantly related to alum treatment, year, and exotic species

frequency of occurrence (Appendix Table 1).

Increases in species richness were observed throughout the study period (Table 2).

The most commonly observed native species throughout all the survey years were

coontail (Ceratophyllum demersum), Canada waterweed (Elodea canadensis), and sago

pondweed (Stuckenia pectinata). Once exotic species were controlled, chara (Chara spp.)

and narrowleaf pondweed (Potamogeton pusillus) were recruited to the lake. Following

38

the increase in water clarity due to the alum treatment, wild celery (Vallisneria

americana) and water stargrass (Heteranthera dubia) were observed in the lake (Figure

3, Table 2). The number of points in August with more than 3 species found at a point

increased from 6 in 2011 to 42 in 2016. Additionally, the maximum depth of rooted

native plant growth increased from 3.1m in 2011 to 4.1m in 2016, although the increase

was not significant (Table 1).

The total native plant community (all native taxa combined) showed large

increases over the study period in both frequency of occurrence and biomass. The total

native plant frequency of occurrence increased from 49% in 2011 to 82% in 2016

(p=0.08) (Figure 4 and Appendix Table 2). The increase occurred steadily over the course

of the survey years as growing conditions improved with exotic species control and water

clarity improvement. The increases in native frequency of occurrence were significantly

related to alum treatment, exotic species frequency, and month. (Appendix Table 2). The

total biomass of all native species increased from 43.8g/m2 ± 11.0g/m2 in 2011 to

707.5g/m2 ± 306.2g/m2 in 2016 (p<0.01) (Figure 5). The native macrophyte biomass was

significantly related to alum treatment, year, month and exotic species biomass (p<0.01)

(Appendix Table 3). The August native plant biomass doubled between 2014 and 2015

and again between 2015 and 2016 (Figure 5). The total native plant biomass was

composed mainly of coontail and Canada waterweed.

Exotic species curlyleaf pondweed and Eurasian watermilfoil also showed

changes over the study period. The curlyleaf pondweed frequency of occurrence was high

after carp removal; in June of 2011 the frequency of occurrence was 34% and in June of

39

2012 it was 28%. Following the herbicide treatments in early spring starting in 2013, the

June curlyleaf pondweed frequency of occurrence remained at 35% or below in 2013

through 2016, the change in frequency was not significant (Figure 3 and Appendix Table

2). Although the June curlyleaf frequency of occurrence was still high, the biomass

decreased greatly after herbicide treatments began in 2013. The June biomass was highest

in 2012 at 120.0g/m2 but never reached greater than 27.4 g/m2 during the years when

treatments occurred (p<0.01) )(Figure 6 and Appendix Table 3). Curlyleaf pondeed

biomass was significantly related to herbicide treatment and month (Appendix Table 3).

The August curlyleaf pondweed frequency of occurrence and biomass was low

throughout the survey years due to the life cycle of the plant in which it peaks its growth

in June and senesces by August each year.

Eurasian watermilfoil frequency of occurrence was high in both June and August

surveys after carp were removed (Figure 3). The treatment of Eurasian watermilfoil was

successful in 2015 and 2016. Each year after treatment in June, the frequency of

occurrence was reduced in August relative to the June frequency (Figure 3). The change

in Eurasian watermilfoil frequency was significant over the study period (p<0.05) and

was significantly related to native frequency of occurrence, year, and month (Appendix

Table 2). Eurasian watermilfoil biomass also decreased in August relative to the June

measurements after herbicide treatments occurred. The average biomass of Eurasian

watermilfoil significantly decreased during the survey years (p<0.01). The Eurasian

watermilfoil biomass was significantly related to native species biomass and month. The

June biomass was highest in 2012 at 135.7 g/m2 and decreased to 52.8 g/m2 in 2015, and

40

11.7 g/m2 in 2016. The August biomass was highest in 2014 at 43.2 g/m2 and was

6.9g/m2 in 2015 and 22.6g/m2 in 2016 after treatments (p<0.05) (Figure 6, Appendix

Table 3).

Coontail was the most commonly occurring native macrophyte throughout the

survey years. The frequency of occurrence increased as growing conditions improved

from 47% in 2011 to 72% in 2016, however the results were not significant (Figure 3,

Appendix Table 2). Coontail biomass increased as a result of the management actions;

the biomass was 41.7g/m2 in 2011 and increased to 673.0g/m2 in 2016 (p<0.05) (Figure

6). Coontail biomass was significantly related to alum treatment, year, month, and the

exotic species biomass (Appendix Table 3). Coontail continued to make up the vast

majority of the native aquatic plant biomass in Lake Riley, with little contribution of

other native plants to total lake-wide native plant biomass.

Canada waterweed was commonly observed during surveys but not densely

growing. The frequency of occurrence increased as growing conditions improved from

2% in 2011 to 38% in 2016 (p<0.01) (Figure 3). The frequency of occurrence was

significantly related to year, month, alum treatment, and exotic species frequency of

occurrence (p<0.05) (Appendix Table 2). The biomass also greatly increased as a result

of the management actions. The biomass was 0.06g/m2 in 2011 and 30.1g/m2 in 2016

(p<0.01) (Figure 6). Similar to the frequency of occurrence analysis, the model for

Canada waterweed biomass indicated that it was significantly related to year, month,

exotic species biomass, and the alum treatment (p<0.01) (Appendix Table 3).

41

Sago pondweed was also regularly observed during surveys. The frequency of

occurrence was stable during the survey period occurring between 0% and 6%

occurrence, there was no significant change in the frequency of occurrence (Figure 3 and

Appendix Table 2). The biomass fluctuated between 0g/m2 and 3.4g/m2 and there was no

significant change (Figure 6 and Appendix Table 3).

Narrowleaf pondweed was also observed during surveys. The frequency of

occurrence fluctuated between 0% and 10% throughout the study (Figure 3). The biomass

fluctuated between 0g/m2 and 0.4g/m2 during the study. There was no significant change

(Figure 6 and Appendix Tables 2 and 3).

Chara was recruited to Lake Riley in 2014. The frequency of occurrence was 0%

for all survey years until 2014 when it was observed at 1% of sites. The frequency of

occurrence in 2015 and 2016 ranged between 1% and 3% (Figure 3). The increase in

frequency of occurrence was significant and was significantly related to exotic species

frequency of occurrence, month, and year (Appendix Table 2). The biomass was 0.0g/m2

until 2015 when it increased to 8.9g/m2, but the biomass decreased to 3.1g/m2 in 2016

(Figure 6). The change was not significant (Appendix Table 3).

Naiad species (Najas flexilis and Najas guadalupensis) also increased in

frequency throughout the survey years. The frequency of occurrence was zero throughout

all survey years until 2015 when it was observed at 6% in 2015 and 2% in 2016 (Figure

3). However, the increase in frequency of occurrence was not significant (Appendix

Table 2). The naiad biomass was 0g/m2 in all survey years until it reached 2.4g/m2 in

2015 and 0.13g/m2 in 2016 (Figure 6). The increase in biomass was significant and

42

significantly related to exotic species biomass, month, and year (p<0.05) (Appendix

Table 3)

Discussion

Overall, the Lake Riley macrophyte community increased in richness and

abundance over the study period. The recruitment of new native species in Lake Riley

was incremental as growing conditions improved throughout the study period. The

comprehensive management on Lake Riley improved the native macrophyte community

and demonstrated the importance of adaptive and multi-year management when

attempting to restore aquatic macrophytes. This study on Lake Riley also emphasizes the

importance of continuous exotic species control when simultaneously working to

improve water clarity. A previous study on the effect of alum treatments in the Twin

Cities metro area showed drastic increases in exotic species when the clarity improved

and no control efforts were in place (Newman et al. 2004). Without the control of the

invasive macrophytes, the alum treatment may have resulted in a dense monoculture of

Eurasian watermilfoil or curlyleaf pondweed due to their rapid growth. However, this

study demonstrated that with consistent, multi-year management of the exotic species

population, the invasive species can be controlled as growing conditions are improved for

native macrophyte species.

Prior to the carp removal, information on the Lake Riley plant community is

sparse. Macrophyte data were gathered by UMN field technicians from 2007 through

2010 (P. G. Bajer and P. W. Sorensen, personal communication). Through a visual

43

assessment, the only vegetation found were intermittent populations of Eurasian

watermilfoil, coontail, curlyleaf pondweed, and Canada waterweed indicating that carp

were limiting the expansion of vegetation in Lake Riley (P. G. Bajer, personal

communication). Additionally, several studies have documented that macrophyte

communities are greatly impaired by large populations of common carp (Bajer et al.

2009, Kloskowski 2011). Therefore, it is likely that if carp reduction did not occur on

Lake Riley the macrophyte community would not have expanded as observed during the

survey years. The July and August water clarity did not improve after carp reduction,

likely due to high levels of phosphorus in the water column resulting in high

phytoplankton abundances (see also Bajer and Sorensen 2015).

After the carp removal, the invasive species curlyleaf pondweed and Eurasian

watermilfoil expanded rapidly in the lake as they can tolerate lower light conditions and

outcompete native species (Bolduan et al. 1994, Chase and Knight 2006). Following

herbicide control, the invasive macrophytes were still present but not at damaging levels.

The curlyleaf pondweed frequency of occurrence and biomass was greatly reduced by the

2016 June survey and the Eurasian watermilfoil biomass was reduced by 2016 although

the frequency of occurrence was still similar to pre-treatment years. After the exotic

species curlyleaf pondweed and Eurasian watermilfoil were controlled, native plants

slightly increased in frequency of occurrence and biomass. However, water clarity was

still limiting the expansion and growth of native plants despite the reduced competition

with the invasive species after herbicide control. After the water clarity improvement in

2016 the exotic species observations remained similar to previous years while native

44

species were observed expanding, suggesting that the herbicide treatments were effective

at controlling the invasives even after growing conditions improved.

After the alum treatment, the native species responded to the improved water

clarity and the species richness further increased. The abundance and biomass also

increased for several native macrophyte species as the growing conditions were enhanced

throughout the survey years. For many species, the largest increases occurred between the

years of 2015 and 2016 when water clarity was drastically improved. Similar to other

studies on macrophyte community recovery, we found that water clarity is a significant

driver in the abundance and diversity of the native plant community (Hilt et al. 2006). In

many cases the observed increases in frequency of occurrence over the survey years were

not significant. However, many species that had insignificant increases in frequency of

occurrence significantly increased in biomass. This pattern may be due to increases in

species density being more pronounced than increases in species expansion throughout

the littoral zone.

Although the native macrophyte community increased in richness and abundance

it was still largely dominated by coontail and Canada waterweed, two native species that

can grow prolifically in the water column and in fact are highly invasive in other regions

of the world (Heikkinen et al. 2009, Hyldgaard and Brix 2012). These taxa were the most

abundant native species by far for both the frequency of occurrence and biomass. The

remaining native macrophyte species all showed moderate to no increases in abundance,

although as light availability increased new species were recruited to the lake during the

survey years. Wild celery and water stargrass were observed in Lake Riley for the first

45

time in 2016; indicative of a healthy plant community, these species are of high value to

the lake ecosystem. Ideally, as water clarity and light conditions remain improved in the

coming years these species will be observed expanding in the lake. Additionally, if the

water clarity and light quality remain at this improved level, transplanting aquatic plants

may be successful in Lake Riley which would further increase the abundance and

diversity.

Overall, to restore a healthy macrophyte community there are several factors that

must be managed in the lake system. Exotic species should be controlled if in high

abundance as the results demonstrated that large populations of exotic species can lead to

stunted native macrophyte populations. Additionally, water clarity appears to be the most

important factor when trying to restore the native plant community. Native plant growth

peaks in August when water clarity tends to be poorest in eutrophic lakes making it

imperative to manage the lake to maintain good water clarity. Long-term planning for

lake management and macrophyte restoration is imperative to ensure successful results

that benefit the ecosystem and lake users.

46

Literature Cited Chapter II

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population of common carp on vegetative cover and waterfowl in a recently

restored Midwestern shallow lake. Hydrobiologia 632: 235-245.

Bajer, P. G. and P. W. Sorensen. 2012. Using Boat Electrofishing to Estimate the

Abundance of Invasive Common Carp in Small Midwestern Lakes. North American

Journal of Fisheries Management 32(5): 817-822.

Bajer, P. G. and P. W. Sorensen. 2015. Effects of common carp on phosphorus

concentrations, water clarity, and vegetation density: a whole system experiment in

a thermally stratified lake. Hydrobiologia 746(1): 303-311.

Bajer, P. G., C. J. Chizinski, and P. W. Sorensen. 2015. Using the Judas technique to

locate and remove wintertime aggregations of invasive common carp. Fisheries

Management and Ecology 18(6): 497-505.

Binzer, T., K. Sand-Jensen, and A. Middelboe. Community photosynthesis of aquatic

macrophytes. Limnology and Oceanography 51(6): 2722-2733.

Bolduan, B. R., G. C. Van Eeckhout, H. W. Quade, and J. E. Gannon. 1994.

Potamogeton crispus–the other invader. Lake and Reservoir Management 10(2):

113-125.

Chase, J. M. and T. M. Knight. 2006. Effects of eutrophication and snails on Eurasian

watermilfoil (Myriophyllum spicatum) invasion. Biological Invasions 8(8): 1643-

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Cooke, G. D., E. B. Welch, S. A. Peterson, and S. A. Nichols. 2016. Restoration and

Management of Lakes and Reservoirs. 3rd ed. Boca Raton, Fla.: CRC, Press.

Cross T. K. and M. C. McInerny. 2006. Relationships between aquatic plant cover and

fish populations based on Minnesota lake survey data. Minnesota Department of

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Dennison, W. C., R. J. Orth, K. A. Moore, J. C. Stevenson, V. Carter, S. Kollar, P. W.

Bergstrom, and R. A. Batiuk. 1993. Assessing water quality with submersed aquatic

vegetation. BioScience 43(2): 86-94.

Getsinger, K. D., D. G. Petty, and J. D. Madsen. 2000. Aquatic dissipation of the

herbicide triclopyr in Lake Minnetonka, Minnesota. Pest Management

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Hilt, S., E. M. Gross, M. Hupfer, H. Morscheid, J. Mählmann, A. Melzer, and K. Van de

Weyer. 2006. Restoration of submerged vegetation in shallow eutrophic lakes – A

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of Inland Waters 36(3): 155-171.

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quality and sediment in the Minneapolis Chain of Lakes, Minnesota, USA. Lake

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Johnson J. A., A. R. Jones, R. M. Newman. 2012. Evaluation of lakewide, early season

herbicide treatments for controlling invasive curlyleaf pondweed (Potamogeton

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Newman, R. M., P. Brezonik, and B. Huser. 2004. Influence of alum treatments on native

and exotic macrophyte density and potential for biological control of Eurasian

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49

Tables Chapter II

Table 1. Summary of maximum depth of plant growth observed, the percent of points

with native submersed taxa, the number of submersed native taxa, and the average Secchi

depth obtained from point-intercept survey data in Lake Riley from 2011 through 2016.

Maximum depth of growth is based on the 95th percentile of points where plants were

observed growing.

Survey Date

Maximum Depth of

Plant Growth

Observed (95%) (m)

% of Points with

Submersed

Native Taxa

Number of

Submersed

Natives

Average

Secchi

Depth (m)

June 2011 4.0 50% 6 4.1

August 2011 3.8 49% 7 0.6

June 2012 4.0 55% 9 2.0

August 2012 3.9 55% 9 0.7

June 2013 3.8 53% 6 2.2

August 2013 3.8 42% 9 0.7

June 2014 3.2 46% 10 1.7

August 2014 3.5 53% 9 2.1

June 2015 3.1 62% 8 1.7

August 2015 3.2 67% 11 1.1

June 2016 4.0 81% 6 3.0

August 2016 4.0 87% 14 1.8

50

Table 2. Aquatic plants found in surveys conducted in Lake Riley 2011 through 2016.

Common Name Scientific Name Abbreviation

Year First

Observed

Submerged species

Coontail Ceratophyllum demersum Cdem 2011

Muskgrass Chara spp. Char 2012

Canada waterweed Elodea canadensis Ecan 2011

Water stargrass Heteranthera dubia Zdub 2016

Bushy Pondweed Najas flexilis Nfle 2011

Southern Naiad Najas guadalupensis Ngua 2015

Northern watermilfoil Myriophyllum sibiricum Msib 2011

Eurasian watermilfoil Myriophyllum spicatum Mspi 2011

Curlyleaf pondweed Potamogeton crispus Pcri 2011

Leafy pondweed Potamogeton foliosus Pfol 2015

Long-leaf pondweed Potamogeton nodosus Pnod 2015

Narrow leaf pondweed Potamogeton pusillus Ppus 2011

Flat-stem pondweed Potamogeton zosteriformis Pzos 2015

Sago pondweed Stuckenia pectinata Spec 2011

Wild celery Valliseneria americana Vame 2016

Horned pondweed Zannichellia palustris Zpal 2011

Floating-leaf Species

Common duckweed Lemna minor Lmin 2014

White lily Nymphaea odorata Nodo 2011

Greater duckweed Spirodela polyrhiza Spol 2012

51

Figures Chapter II

Figure 1. Riley Creek watershed with Lake Riley highlighted in yellow (source: Riley

Purgatory Bluff Creek Watershed District, Chanhassen, MN).

52

Figure 2. Lake Riley average Secchi depth (m) throughout the study period in spring

(May), early summer (June), and late summer (August).

0

1

2

3

4

5

6

7

2011 2012 2013 2014 2015 2016

Sec

chi

Dep

th (

m)

May

June

August

53

Figure 3. The frequency of occurrence of the most common species in Lake Riley in June

and August surveys from 2011 to 2016. EWM stands for Eurasian watermilfoil, CLP

stands for curlyleaf pondweed.

0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

100%F

req

uen

cy o

f O

ccu

rren

ce

11-Jun

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15-Jun

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0%

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30%

40%

50%

60%

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100%

Fre

quen

cy o

f O

ccurr

ence

11-Aug

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14-Aug

15-Aug

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54

Figure 4. The frequency of occurrence for all native species combined and all exotic

species combined in June and August surveys in the years of 2011 through 2016.

Figure 5. The biomass for all native species combined and all exotic species combined in

the June and August surveys in the years of 2011 through 2016. Data are displayed on a

log scale.

1

10

100

1000

Dry

Pla

nt

Bio

mas

s (g

dry

/m2)

±2S

E

Native Exotic Native Exotic

201120122013201420152016

June August

55

Figure 6. The biomass of the most commonly observed species in Lake Riley in the June

and August surveys from 2011 through 2016. Species abbreviations are located in Table

2. Data are displayed on a log scale.

1

50

Cdem Chara Ecan Mspi Niad Nodo Pcri Pfol Ppus Pzos Spec Vame Zdub

Dry

Pla

nt

Bio

mas

s (g

dry

/m2)

±2

SE

11-Jun

12-Jun

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Cdem Chara Ecan Mspi Niad Nodo Pcri Pfol Ppus Pzos Spec Vame ZdubDry

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/m2)

±2

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11-Aug

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56

Chapter III

The Response of Macrophyte Propagules to Light Quantity and the Comparison of

Viable Lake Seed Banks to the Existing Macrophyte Community

57

Summary

The importance of the aquatic macrophyte seed bank to sustaining macrophyte

communities in lakes is an area in need of study. Although vegetative propagation is

common, the seed bank may be contributing more to macrophyte recruitment than what is

currently understood. This is potentially the case in eutrophic lakes that have had poor

water clarity and few macrophytes for several years and then undergo management to

improve clarity. Methods to enhance native macrophytes after improved water quality are

needed. Moreover, the role of water clarity in recruitment from the seed bank is not

understood.

I aimed to understand the role of the seed bank in the recruitment of macrophytes

after water clarity improvement. A controlled laboratory experiment was conducted using

sediment from Lakes Ann and Riley located in Chanhassen, MN, to 1) assess the

germination response under different treatments and 2) compare the observed sprouting

taxa to the taxa growing in the lakes. The treatments included a maximum germination

treatment, a treatment representative of a lake with good water clarity, and a treatment

representative of a lake with poor water clarity.

The observed viable seed banks of both Lakes Riley and Ann reflected the

macrophyte community actively growing in the lake. In Lake Ann, every species

observed in the experiment was observed growing in the lake. In Lake Riley, all but two

species, Richardson’s pondweed and Robbins’ pondweed, observed in the seed bank

were also found growing in the lake. The species observed in the seed bank were not,

however, similar to the observed relative abundance in the lakes.

58

The seed banks did not show any significant difference in response to the

germination treatments. The most frequent species observed in the seed banks were

chara, curlyleaf pondweed, and wild celery. Seventeen species were observed in Lake

Riley and 16 in Lake Ann. Under maximum germination conditions, Lake Riley had a

viable vascular seed density of 2,916 ± 1,828 seeds/m2 and a viable chara spore density

of 1,033 ± 698 spores/m2. Lake Ann had a viable vascular seed density of 1,100 ± 440

seeds/m2 and viable chara spore density of 13,833 ± 2,825 spores/m2. Overall, the study

demonstrated that germinating propagules from a lake seed bank can be a valuable tool

for managers to evaluate the viable macrophyte taxa present and the potential for

recruitment from the seed bank.

Introduction

Macrophytes play an integral role in aquatic littoral zones. Macrophytes stabilize

sediment and sequester nutrients, maintaining water clarity and reducing the potential for

harmful algal blooms and other water quality impairments (Dennison et al. 1993,

Horppila and Nurminen 2003, Bakker et al. 2010). Diverse, heterogeneous aquatic plant

communities also provide habitat for invertebrates and fish communities in a lake (Valley

et al. 2004, Cross and McInerny 2006). When large scale disturbances occur in a lake that

affect macrophytes, such as benthivorous fish damage or high nutrient levels, the

macrophyte community is often in low abundance and primarily a dense monoculture of

an invasive plant due to poor water clarity and growing conditions (Chase and Knight

2006, Bajer et al. 2009). Subsequently, the reduced macrophyte population further

59

impairs the functioning of the ecosystem and the water quality as they are no longer

sequestering nutrients and sediment or providing habitat for other aquatic organisms

(Scheffer et al. 1993, Hansson et al. 1998). As a result, these systems have poor

recreational value for lake users and provide poor habitat for aquatic organisms.

Therefore, the management and restoration of aquatic macrophytes is often a

primary goal when aiming to improve water quality and clarity in a lake (Scheffer et al.

1993). The restoration of native macrophyte communities is a dynamic process that

regularly requires several steps and multiple years of active management. There are

several disturbance factors that can be managed when attempting to improve the growing

conditions for native macrophytes. In Minnesota lakes, eutrophication and invasive fish

and macrophytes can reduce the abundance and diversity of native aquatic plant

communities. Benthivorous fish, mainly common carp (Cyprinus carpio), disturb the

sediment and uproot macrophytes during feeding and spawning (Bajer et al. 2009).

Invasive macrophytes, such as curlyleaf pondweed (Potamogeton crispus) and Eurasian

watermilfoil (Myriophyllum spicatum), can outcompete native species and subsequently

reduce the diversity of the macrophyte community (Madsen et al. 1991, Johnson et. al

2012, Jones et al. 2012). Additionally, excess nutrients, including phosphorus and

nitrogen, can cause the proliferation of planktonic algae and reduce water clarity, shading

out macrophytes (Scheffer et al. 1993). Therefore, restorative actions often include

nutrient reduction, nutrient sequestration, and invasive species control (Cooke et al.

2016). These mechanisms aim to improve the growing conditions for native macrophytes

60

by decreasing competition with non-native macrophytes and by increasing the water

clarity and subsequently the light availability allowing for efficient photosynthesis.

Often, when water clarity is improved through management actions, such as a

benthivorous fish removal or alum treatment, the native macrophyte community increases

in abundance and diversity. Specifically, in Midwestern lakes when common carp were

reduced to a biomass of less than 100kg/hectare, the aquatic plant community was

documented to improve (Bajer et al. 2009, Bajer and Sorensen 2015). The reduction of

carp allows aquatic plant communities to improve due to the reduced sediment

disturbance and improved water clarity that can occur when carp are reduced. Also,

increases in macrophyte communities have been documented after water clarity improves

due to alum or bentonite treatment (Spears et al. 2016). However, in some cases increases

in macrophyte abundance can be largely due to an invasive species, such as Eurasian

watermilfoil (Myriophyllum spicatum) (Jacocby et al. 1994, Newman et al. 2004).

Therefore, actions to improve water clarity are often paired with invasive macrophyte

control measures to allow for recovery of a diverse, heterogeneous native plant

community, which improves overall lake ecosystem functioning (Hilt et al. 2006, Cooke

et al. 2016).

When management of invasive fish and/or macrophytes occurs without an

increase in water clarity, the native macrophyte community may show a marginal

increase in abundance and species diversity (Knopik 2014, JaKa 2015). Therefore, it

appears that to restore native macrophyte communities, water clarity must be maintained

at a high enough level throughout the summer growing season to facilitate the

61

recruitment of native macrophyte species to the lake (Chambers and Kalff 1985, Doyle

and Smart 2001, Knopik 2014).

The recruitment of new species has been documented in lakes that have

undergone a water clarity improvement. However, it is unknown if the recruitment is due

to growth from existing source populations or the seed bank. Native macrophytes mainly

propagate through clonal growth and fragmentation (Santamaría 2002). Many species can

send out roots from which a new vegetative structure can sprout or they can grow a new

plant from small fragments of a plant. Although most aquatic macrophytes propagate

mainly through clonal growth they do still produce seeds or spores (Santamaría 2002,

Boedeltje et al. 2003). However, the propagules that are produced are generally not

thought be a large contributor to macrophyte community propagation and recruitment

(Boedeltje et al. 2003).

Although propagation through clonal growth is most common, in some systems

the source populations of submersed aquatic vegetation may be absent due to low clarity

or benthivorous fish disturbance. This absence of a macrophyte community may lead to

the role of the seed bank being more influential in the restoration of lake vegetation (De

Winton et al. 2000, Pollux 2011). Overall, there are few studies that have assessed the

relative role of the submersed macrophyte seed bank in the revegetation of a lake.

Moreover, there is limited understanding on the extent that species rely on sexual and

asexual modes of reproduction. Initial studies on submersed macrophyte species have

found varying levels of asexual and sexual reproduction. For instance, Najas minor has

been demonstrated to propagate mainly through seeds (Les et al. 2015), whereas Eurasian

62

watermilfoil and water hyacynith have been demonstrated to rely more on propagation

through clonal growth, and have low genetic diversity across their native range (Wu et al.

2015, Zhang et al. 2010). Overall, asexual versus sexual reproduction appears to vary by

species and environmental conditions (Pollux et al. 2007). Specifically regarding sexual

reproduction, several studies have evaluated the germination of submersed aquatic plant

propagules but these studies have been focused on one or two species and not on the seed

bank of a lake as a whole (Hartleb 1993, Jarvis and Moore 2008, Xiao 2010).

Additionally, the majority of aquatic plant seed bank research has been conducted on

emergent wetland species, which have different germination requirements than

submersed aquatic vegetation (Baskin and Baskin 2014). These studies have found that

light is an important factor in the germination process for some species (Coble and Vance

1987, Titus and Hoover 1991, Dugdale et al. 2001, Baskin and Baskin 2014).

My study aimed to assess the effect of improved water clarity and improved light

quantity on the germination of macrophyte propagules from lake seed banks.

Specifically, sediment was placed in one of three treatments, each with different

environmental conditions, to evaluate the effect of water clarity on sprouting. The total

number of sprouted propagules as well as number of sprouted propagules per species

were counted to obtain an estimate of viable seed density. Propagule sprout counts were

also separated into the total vascular seeds counted and the total macroalgae spores

counted. Subsamples of sediment were also enumerated to estimate the number of seeds

that did not germinate.

63

The aim of this study is twofold. Firstly, we were interested in estimating the seed

bank in two Minnesota lakes, one with a diverse plant community and the other with a

historically limited plant community due to low clarity and carp disturbance, to

understand if the seed bank reflects the current occurrence of macrophytes in a lake.

Secondly, we wanted to determine the role of light intensity on the germination of

macrophyte propagules. Understanding how the seed bank of a lake responds to lake

restoration actions, such as a water clarity/light intensity improvement, is key for a

comprehensive macrophyte restoration project. Having an insight into what triggers

germination from the seed bank is vital as lake managers plan for desired outcomes, such

as a more stable, diverse macrophyte community.

Methods

Seed Bank Collection

Sediment was collected from Lake Ann (DOW ID 10001200) and Lake Riley

(DOW ID 10000200). Both lakes are in Chanhassen, MN, in Carver County, west of the

Twin Cities within the Riley Purgatory Bluff Creek Watershed District (Figure 1). Lake

Ann is a dimictic, mesotrophic lake with a diverse macrophyte community and good

water clarity (mean August Secchi depth: 1.8m). In Lake Ann, 17 to 21 species have been

regularly observed in 2011 to 2014. The lake area is 48 hectares with a maximum depth

of 12.2m. Lake Riley is a eutrophic lake with a historically diminished macrophyte

population that has been steadily improving due to lake management actions over the last

several years (carp reduction, invasive macrophyte control, and alum treatment). The lake

64

area is 120 hectares with a maximum depth of 15.1m. The mean August Secchi depth

was 0.5m prior to the 2016 alum treatment. In Lake Riley, 12-15 species have been

regularly observed between the years of 2014 to 2016, previously 7 to 10 species were

regularly observed in 2011 to 2013.

Sediment was collected at Lake Riley on May 5th and 6th, 2016 and at Lake Ann

on May 19th, 2016. To collect the sediment for the treatments, seven transects were

marked around each lake using ArcGIS. Transects were uploaded to a Garmin 76 GPS

device and a boat was navigated to each site. At each transect, four sediment core

samples were obtained at a 1.0m depth using a 10.0cm diameter PVC coring device; the

top 5.0cm of the sediment core was collected. The sediment samples from each lake were

homogenized to reduce the heterogeneity of seed bank distribution in lake and then stored

in a dark refrigerator at 4 ˚C until the experiment was ready to begin.

Seed Bank Treatments

After sediment collection, the sediment from Lakes Ann and Riley was allocated

into small trays. Sediment samples were washed with well water over a coarse sieve to

remove large material, such as twigs and cobbles, and vegetative structures and buds,

such as curlyleaf pondweed turions. After the material was removed, 200mL of sediment

was spread in a layer over a medium of 200mL sterilized sand in 19.0cm x 19.0cm x

6.0cm trays (Galatowitsch 1994) and covered with 3.0cm of water (Boedeltje 2002,

Baskin and Baskin 2014). A total of 45 trays were created for each lake. Fifteen

additional trays were used as controls to ensure that contamination in the growing room

65

did not occur. Control trays consisted of 200mL steam sterilized sand and 200mL

sterilized lake sediment.

Trays were allocated to one of three germination treatments of varying light

intensity and the experiment ran for sixteen weeks. Trays were illuminated with

Helioscpectra lights. Heliospectra lights emit nine wavelengths of light that can each be

individually adjusted to precisely control the light. The lights have 380nm, 400nm,

420nm, 450nm, 520nm, 630nm, 660nm, and 735nm wavelengths in addition to 5700K

white LED lights that are similar to sunlight. Curtains were placed between different light

treatments to eliminate the effect of other light sources on the propagules.

To assess the extent of the viable seed bank in Lakes Ann and Riley, 15 of the

trays for each lake were used to assess the viability using the seedling emergence method

to maximize germination (Boedeltje 2002). To maximize germination, the trays were

exposed to a series of environmental conditions known to induce germination in aquatic

plants. The photosynthetically active radiation (PAR) was set to 800 µmol/m2/s and a 15

hour light, 9 hour dark photoperiod (Coble and Vance 1987, Boedeltje 2002). All the

wavelengths of light and white light were kept at equal intensity set to maintain the 800

µmol/s/m2 PAR. The temperature ranged between 21°C and 23°C (Boedeltje 2002), and

water levels remained at approximately 3.0cm in the trays throughout the testing period

(Boedeltje 2002, Baskin and Baskin 2014). Gibberellic acid, to induce sprouting, was

applied to the trays once at the onset of the experiment to reach a concentration of 0.3mM

(Tuckett et al. 2010, Baskin and Baskin 2014).

66

In addition to assessing the total viability of the seed banks in Lakes Ann and

Riley, the effect of water clarity and light quantity on submersed aquatic vegetation

germination in lake seed banks was assessed by exposing the seed bank to one of two

levels of light intensity. Fifteen trays were exposed to a light condition representative of a

clear lake at 1.0m depth and the remaining fifteen trays were exposed to a light condition

representative of a turbid lake at 1.0m depth. For this experiment, a “clear lake” was

defined as having an August Secchi depth of 1.5m or greater and a “turbid lake” was

defined as an August Secchi depth of less than 1.5m.

Field observations with a spectroradiometer were collected in 6 lakes of varying

clarity in the Riley Purgatory Bluff Creek Watershed District at mid-day on a day with no

cloud cover and little wind. These observations indicated that high clarity lakes had an

average light intensity of 650 µmol/s/m2 at 1.0m depth and low clarity lakes 125

µmol/s/m2 at 1.0m. The light spectrum observed at 1.0m was very similar among the

lakes and therefore the wavelengths remained at the same ratios for each treatment and

only the overall intensity of the light was altered. Thus, in the high clarity treatment the

Heliospectra lights were set at a PAR of 650µmol/s with an equal intensity of all nine

wavelengths. In the low clarity treatment the lights were set at 125µmol/s PAR with an

equal intensity of all nine wavelengths. For both treatments a 15 hour light: 9 hour dark

photo period was used. In the good and low clarity treatments, the water temperature was

consistently between 21.0°C to 23.0°C in the trays and the dissolved oxygen was

consistently between 7.0 and 9.0 mg/L.

67

Trays were checked weekly and new propagule sprouts were identified to species

and recorded. Any observed sprouting from the sediment was counted as germinated and

was considered viable (Boedeltje 2002). Propagule sprouts were removed from the trays

after being counted to prevent counting multiple times. If needed, propagules were

transplanted into an environmental chamber for continued growth to confirm species

identification. When trays were checked for sprouting, temperature and dissolved

oxygen readings were made using a YSI ODO sensor every week at midday when the

lights had been on for at least 5 hours.

After the experiment was concluded, a subsample of five trays from each

treatment were examined to enumerate the seed bank to evaluate the number and species

of propagules that did not germinate (Bernhardt et al. 2008). Sediment was sifted through

1.0mm, 0.5mm, 0.25mm, 0.125mm, and 0.053mm sieves stacked on each other to sort

the sample by grain size and more easily find all ungerminated propagules. The sediment

in each sieve was visually inspected and propagules were picked and identified to genus

or species (depending on the morphological characteristics) using a Nikon stereo

microscope.

Lastly, the viable seed bank was compared to the plants observed growing in the

lake through point-intercept surveys to evaluate if the seed bank is representative of the

existing plant community. The species observed sprouting were ranked by most abundant

(rank=1) to least abundant and the species observed during the surveys were ranked by

most abundant (rank=1) to least. These values were plotted on a scatterplot for each lake

68

to evaluate if the most abundant sprouts were also the most abundant plants observed in

the lake.

Sediment Analysis

Part of the homogenized sediment from Lakes Ann and Riley was used for

analysis of the soil characteristics, specifically dry bulk density and organic matter

content. To determine the dry bulk density and organic matter content, five 10.0mL

subsamples of sediment from each lake were obtained using a modified syringe. The

subsample was placed in a crucible that was weighed and recorded prior to the sediment

being added. The subsamples were dried in an oven at 100°C for 48 hours. After drying,

the samples were weighed and the dry bulk density was calculated as g dry/mL. The

samples were then placed in a muffle furnace for 3 hours at 500°C to combust the organic

content in the sediment. The samples were cooled and promptly weighed. The organic

matter content was obtained by subtracting the mass of the sediment after the muffle

furnace from the mass of the sediment after the drying oven; percent organic matter was

then calculated as a percentage of the dry sediment mass. The five subsamples of dry

bulk density and organic matter content for each lake were averaged to obtain a mean dry

bulk density and organic matter content for the sediment.

Statistical Analysis

The study compared the germination of seeds from the seedling emergence

method to the abundance and richness of propagule sprouts in the two light treatments. I

also compared the viable seed bank to the observed species growing in the lakes. The

69

Lakes Riley and Ann seed banks were assessed for several metrics. The number of total

propagule sprouts, number of propagules per species, number of species germinated, the

average total viable seeds per tray and per square meter lake bottom, and the average

viable seeds per tray and per square meter for each species was calculated. Additionally,

the metrics of average viable propagule density per tray and per square meter was

separated into vascular plants that produce seeds and the non-vascular plant spores from

Chara spp. to better understand the composition of the lake seed banks as chara was a

large contributor to the total propagule count in each lake. The average viable propagules

per square meter of lake bottom was determined by taking the average of the count of

propagules in each tray and dividing by 40cm2 to obtain the average propagules per cm2.

The sediment core volume (10.0cm diameter and 5.0cm depth) was 395cm3 representing

a surface area of 79cm2. Thus the 200cm3 of sediment in each tray represents a surface

area of 40cm2. The propagules/cm2 value was then multiplied by 10,000 to obtain the

propagules/m2.

All statistical analysis was conducted using R statistical software version 3.3.2

(The R Foundation for Statistical Computing, 2016). Results were considered statistically

significant when p values were < 0.05. To assess the effect of the treatment on the

number of propagules counted for both Lake Ann and Lake Riley samples, Poisson log-

linear models were used as the data followed Poisson distributions. The models were

used to evaluate the total count of propagules as well as the count of each species

observed germinating. A Poisson log-linear model was also used to evaluate the

significance of the difference between the species richness observed in each treatment.

70

Results

Lake Riley

Overall, in the Lake Riley samples the maximum germination and high clarity

treatments had 13 taxa observed sprouting and the low clarity treatment had 10 taxa

observed sprouting (Table 1). These differences were not significant (p > 0.05). The total

number of propagules germinated was approached by week 8 and was reached by week

16 (Figure 2).

Under maximum germination conditions, the average number of propagules

germinated per tray at the end of the experiment was 15.9 ± 6.24 propagules/tray (Figure

2). Therefore, the mean number of viable propagules was 3,950 ± 1,561 propagules/m2.

The mean number of vascular seeds germinated per tray was 11.6 ± 7.3 seeds/tray or

2,916 ± 1,828 seeds/m2. The mean number of chara spores germinated per tray was 4.1 ±

2.8 spores/tray or 1,033 ± 698 spores/m2 (Figure 3). In the maximum germination

conditions, curlyleaf pondweed had the greatest number of viable seeds observed with an

average of 6.3 ± 3.1 seeds/tray or 1,583 ± 776 seeds/m2 (Figure 3, Table 1). We are

confident that these were from seeds because all visible turions had been removed from

the sediment. Chara was the second most common propagule (Figure 3, Table 1). Wild

celery also had a high number of seeds, at an average of 2.6 ± 1.6 seeds/tray or 650 ± 399

seeds/m2 (Figure 3, Table 1).

Relative to the maximum germination treatment, the Lake Riley sediment samples

exposed to high clarity and low clarity treatments had a lower level of germination, but

the differences were not significant. For the high clarity treatment, the average

71

germination by 16 weeks was 9.3 ± 1.7 propagules/tray or 1,783 ± 444 propagules/m2

(Figure 2). The mean number of vascular seeds germinated per tray was 4.5 ± 1.3

seeds/tray or 1,116 ± 312 seeds/m2. The mean number of chara spores germinated per

tray was 2.7 ± 1.4 spores/tray or 667 ± 341 spores/m2. In the high clarity condition, chara

was the most frequent species observed (Figure 3, Table 1). The second most abundant

species was curlyleaf pondweed with an average germination of 1.7 ± 0.63 seeds/tray or

433 ± 158 seeds/m2 (Figure 3, Table 1). Wild celery was also observed at an average 1.3

± 0.6 seeds/tray or 333 ± 152 seeds/m2 (Figure 3, Table 1). The remaining species were

all observed in low frequencies.

The Lake Riley sediment samples exposed to the low clarity treatment had an

average germination per tray of 7.1 ± 2.7 propagules/tray or 2,167 ± 683 propagules/m2

(Figure 2). The mean number of vascular seeds germinated per tray was 5.4 ± 2.6

seeds/tray or 1,350 ± 653 seeds/m2. The mean number of chara spores germinated per

tray was 3.3 ± 2.3 spores/tray or 817 ± 574 spores/m2. In the low clarity conditions, chara

was the most abundant observed at an average of 3.3 ± 2.3 spores/tray or 816 ± 574

spores/m2 (Figure 3, Table 1). Curlyleaf was also abundant with an average of 3.1 ± 2.1

seeds/tray or 767 ± 533 seeds/m2 (Figure 2, Table 1). Wild celery was also abundant with

an average of 1.47 ± 0.84 seeds/tray or 367 ± 212 seeds/m2 (Figure 2, Table 1). The

remaining species observed were in low abundance.

Overall, the results of the Poisson model indicated no significant difference

between the different treatment types for the Lake Riley samples. Specifically, there was

no difference in the total number of propagules, species diversity, or species abundance

72

(all p > 0.05) (Appendix Table 4). All taxa were that were observed germinating in the

maximum germination conditions, apart from two species (Potamogeton richardsonii and

Potamogeton robinsii), were also observed growing in the lake during macrophyte

surveys. Two taxa that were not present in previous years appeared in Lake Riley during

the summer of 2016, wild celery (Vallisneria americana) and water stargrass

(Heteranthera dubia).

The Lake Riley seed bank enumeration yielded few additional propagules that had

not germinated in the maximum germination conditions. The maximum germination

conditions yielded the fewest propagules that had not germinated while the good and low

clarity conditions had a slightly higher abundance of propagules observed in each sub

sample (Table 2). An average of 1.3 ± 0.42 propagules/tray for the maximum germination

condition was counted. The good and low clarity treatments had a slightly higher number

of remaining seeds with an average of 5.0 ± 2.3 propagules/tray and 6.1 ± 2.5

propagules/tray respectively. No species were found as propagules that had not also been

observed germinating.

The average dry bulk density of the collected Lake Riley sediment was 0.53 ±

0.34 g/mL and the average organic matter content was 13% ± 4.5%.

Lake Ann

Lake Ann had a total propagule count that was much greater than Lake Riley,

largely due to a high occurrence of chara spores. In the maximum germination

conditions, 13 taxa were observed, in the good and low clarity treatments 10 taxa were

73

observed. The maximum germination was approached by week 12 and was complete by

week 16 in Lake Ann.

Under maximum germination conditions the average total germination after 16

weeks was 59.7 ± 11.1 propagules/tray or 14,933 ± 2,771 propagules/m2 (Figure 4). The

mean number of vascular seeds germinated per tray was 4.4 ± 1.6 seeds/tray or 1,100 ±

408 seeds/m2. The mean number of chara spores germinated per tray was 55.3 ± 11.3

spores/tray or 13,833 ± 2,825 spores/m2 (Figure 5, Table 3). The second most abundant

species was wild celery, observed at an average of 2.1 ± 1.2 seeds/tray or 533 ± 296

seeds/m2 (Figure 5, Table 3). Curlyleaf pondweed was observed at an average of 0.9 ±

0.45 seeds/tray or 233 ± 142 seeds/m2 (Figure 5, Table 3). The remaining species

observed were all at low abundances.

The Lake Ann sediment samples exposed to high clarity and low clarity

conditions had a similar level of germination relative to the maximum germination

conditions; the difference between treatments was not significant (p>0.05) For the high

clarity treatment, the average germination per tray in each treatment was 58.6 ± 13.4

propagules/tray or 14,000 ± 3,351 propagules/m2 (Figure 4). The mean number of

vascular seeds germinated per tray was 4.5 ± 1.8 seeds/tray or 1,125 ± 446 seeds/m2. The

mean number of chara spores germinated per tray was 53.4 ± 12.6 spores/tray or 13,350 ±

3,149 spores/m2 (Figure 5, Table 3). Chara was the most abundant taxa. The second most

abundant species was wild celery with an average of 2.5 ± 1.1 seeds/tray or 633 ± 275

seeds/m2 (Figure 5, Table 3). Curlyleaf was also present at high levels at an average of

74

1.3 ± 0.8 seeds/tray or 333 ± 193 seeds/m2 (Figure 5, Table 3). The remaining species

were all in low abundances.

For low clarity conditions, the total was an average of 58.1 ± 17.1

propagules/tray. The mean number of germinated propagules was 16,286 ± 4266

propagules/m2 in the low clarity treatment (Figure 4). The mean number of vascular seeds

germinated per tray was 4.0 ± 1.7 seeds/tray or 1,000 ± 428 seeds/m2. The mean number

of chara spores germinated per tray was 61.1 ± 16.9 spores/tray or 15,286 ± 4,218

spores/m2. The most abundant species was chara (Figure 5, Table 3). The second most

abundant species was wild celery 1.9 ± 0.8 seeds/tray with an average of 482 ± 210

seeds/m2 (Figure 5, Table 3). Curlyleaf was also abundant with an average of 0.8 ± 0.1

seeds/tray or 286 ± 196 seeds/m2 (Figure 5, Table 3).

The results of the Poisson model indicated no significant difference between the

different treatment types for the Lake Ann samples. Specifically, there was no difference

in the total number of sprouts, species diversity, or species abundance (Appendix Table

4). All taxa were that were observed germinating in the maximum germination conditions

were also observed growing in the lake during macrophyte surveys.

The Lake Ann seed bank enumeration yielded few additional propagules that had

not germinated. An average of 2.3 ± 1.4 propagules/tray remained from the maximum

germination treatment (Table 2). An average of 1.3 ± 1.0 propagules/tray remained from

the high clarity condition (Table 2). An average of 2.3 ± 1.7 propagules/tray were

counted from the low clarity condition (Table 2). No species were found as propagules

that had not also been observed as sprouts. Lastly, the average dry bulk density of the

75

Lake Ann sediment was 0.72 ± 0.12 g/mL and the average organic matter content was

21% ± 5%.

Overall, the seed banks of Lakes Ann and Riley were distinct despite being within

the same watershed. The Lake Ann seed bank had a higher number of chara spores in all

treatments relative to the Lake Riley seed bank and the difference was significant

(p<0.05, t-test). The vascular seed count was similar among the two lakes for each

treatment and there was no significant difference in the vascular seed counts between the

two lakes or among the three treatments.

Discussion

Although germination was higher in the maximum germination treatment for both

lakes, there was high variability and no significant increase in germination with that

treatment for either lake relative to the good and low clarity treatments (p>0.5). Despite

the findings in other studies (Dugdale et al. 2001, Sederias and Colman 2007), in this

study light quantity did not appear to have a significant effect on the germination of

propagules from a lake seed bank. Other studies have suggested that temperature and

burial depth in the sediment are also key factors in dormancy breaking and germination

(Baskin and Baskin 2014). In this experiment, these other factors may have been more

critical to germination than light.

Interestingly, the timing of the germination was variable between the two lakes

despite having similar compositions of taxa present (Figures 2, 4). Germination was

observed in both lake sediments beginning at week two. However, the germination in the

Lake Riley sediments leveled off roughly after 8 weeks whereas in Lake Ann

76

germination continued through week 12. Overall, Lake Ann had a much higher viable

propagule count in all treatments relative to Lake Riley, due to the many chara spores in

the sediment in Lake Ann (Figure 5). However, there was no difference in the number or

species of vascular seeds counted in the trays between both lakes despite Lake Ann

having a greater level of macrophyte diversity. This study demonstrates that the seed

banks of lakes, even within the same drainage, can be variable regarding the abundance

of propagules such as is the case with chara.

These results provide important information that will guide future studies on

macrophyte seed banks. Light intensity did not have an effect on germination, however it

is likely an important factor promoting propagule growth and development into a mature

plant (Jarvis and Moore 2008). Further evaluation of seedling survival in different light

conditions is needed, such as evaluating responses in lower PAR conditions such as 25

µmol/s where the light may be under the compensation point for the plant.

Assessment of the seed bank also provides useful information to lake managers

regarding invasive species management. Prior to the sediment collection in 2016, in Lake

Riley, curlyleaf pondweed was treated with endothall herbicide for three consecutive

years in May of 2014, 2015, and 2016 and the control efforts were successful at reducing

the abundance of curlyleaf pondweed growth in Lake Riley. The fall turion densities in

the sediment declined significantly from 61 ± 20 turions/m2 in 2012 to to 2 ± 1.4

turions/m2 in 2015 (p<0.05) (Dunne and Newman 2017). Suppression or depletion of

turions is an important strategy for curlyleaf pondweed control but it is difficult to

achieve (Crowell and Madsen 1988, Johnson et al. 2012). In the Lake Riley seed bank

77

samples, curlyleaf pondweed was consistently one of the most abundant species

sprouting, with an estimated viable seed density at 175 seeds/m2 in the sediment as

compared to the viable turion density of 2 turions/m2 in 2016. There is clearly still an

abundant and viable propagule source that managers will need to be aware of as they

manage this plant over the next several years as recruitment from seed may be more

common than generally thought.

Eurasian watermilfoil also occurs in Lakes Ann and Riley in relatively high

abundances at certain locations in the lakes. However, this species had a low abundance

in the seed bank based on sediment samples in both lakes. This finding is consistent with

other studies on Eurasian watermilfoil propagation indicating that this species may rely

mainly on fragmentation and clonal growth for its propagation (Coble and Vance 1987,

Madsen and Smith 1997). However, viable seeds were present and did sprout in Lake

Riley sediments at low levels.

In addition to aiding the understanding of invasive species populations, by

employing a germination study, lake managers can also better understand the potential

for native species recolonization from the seed bank in a lake. Specifically, lake

managers can determine the extent of species diversity present in the lake and what taxa

have the potential to recolonize. For example, in Lake Riley two species were observed

sprouting in this experiment that had not been observed in Lake Riley during aquatic

vegetation point-intercept surveys that occurred from 2011 to 2016. Richardson’s

pondweed (Potamogeton richardsonii) and Robbins’ pondweed (Potamogeton robinsii)

were observed as sprouts and this indicates that there may be the potential for recruitment

78

from the seed bank in Lake Riley. Additionally, several species that were observed in the

seed bank were only observed growing in Lake Riley after invasive species management

and water clarity improvement, including floating leaf (Potamogeton nodosus) and flat

stem pondweeds (Potamogeton zosteriformis) in 2015, and water stargrass (Heteranthera

dubia) and wild celery (Vallisneria americana) in 2016. This study suggests that those

species may have been recruited from seed due to the improvement in growing

conditions. In Lake Ann, there was a high diversity of taxa observed during the 2011

through 2014 survey years. In the Lake Ann seed bank all species observed sprouting

were also observed during point-intercept surveys. The scatterplots evaluating the

relationship between seed bank abundance and lake abundance showed no relationship. If

there were a relationship between abundance in the seed bank and the lake, one would

expect the points to linearly align with a slope of one, however this is not the case in

either lake. There was no clear pattern of sprout density and observed plant density

indicating that high abundance in the seed bank does not equate to high abundance in the

lake (Figures 6 and 7).

Overall, lake seed banks can be variable in abundance and richness and in this

study the seed banks appear to reflect the existing macrophyte community in the lakes.

Moreover, water clarity and ranges of high light intensity did not impact the propagule

germination of macrophytes in our study but further investigation is warranted as to the

effect of light on the survival of propagules to maturity.

79

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84

Tables Chapter III

Table 1. Total number of propagules germinating from Lake Riley sediment for the

maximum germination, high clarity, and low clarity treatments based on 3.0L of collected

sediment per treatment. Treatment Type

Species Species

Abbreviation

Maximum

Germination

High

clarity

Low

clarity

Ceratophyllum demersum Cdem 10 8 3

Chara spp. Char 62 40 49

Elodea canadensis Ecan 3 5 1

Heteranthera dubia Hdub 1 1 0

Lemna minor Lmin 1 2 0

Lemna trisulca Ltri 0 0 0

Myriophyllum spicatum Mspi 1 1 0

Najas guadalupensis Ngua 1 0 0

Nyphar varigaeta Nvar 1 0 0

Potamogeton crispus Pcri 95 26 46

Potamogeton pusillus Ppus 13 8 5

Potamogeton nodosus Pnod 9 1 3

Potamogeton robinsii Prob 1 1 0

Potamogeton zosteriformis Pzos 4 3 3

Ranunculus longirostris Rlon 9 0 0

Stuckenia pectinata Spec 39 4 1

Vallisneria americana Vame 3 20 22

Table 2. Total ungerminated seeds enumerated from the subsample of five Lake Riley

and five Ann trays. Lake Riley Lake Ann

Species Max.

Germination

High

clarity

Low

clarity

Max.

Germination

High

clarity

Low

clarity

Mspi 1 0 0 3 0 2

Niad 1 2 2 1 2 0

Pamp 0 1 0 0 0 4

Pcri 0 2 1 0 2 1

Ppus 2 0 1 0 0 0

Prob 0 1 0 1 0 0

Pzos 0 0 1 2 0 0

Spec 0 0 0 0 0 0

Zpal 0 2 3 0 0 0

85

Table 3. Total germination of propagules from Lake Ann sediment for the maximum

germination, high clarity, and low clarity treatments based on 3.0L of collected sediment

per treatment.

Treatment Type

Species Species

Abbreviation

Maximum

Germination

High

clarity

Low

clarity

Ceratophyllum demersum Cdem 1 2 3

Chara spp. Char 830 801 911

Lemna minor Lmin 0 2 0

Lemna trisulca Ltri 2 0 0

Najas guadalupensis Ngua 2 0 0

Nutela lutembo Nlut 1 0 0

Nyphar varigaeta Nvar 1 0 0

Potamogeton crispus Pcri 14 20 18

Potamogeton pusillus Ppus 5 5 11

Potamogeton nodosus Pnod 1 0 1

Potamogeton richardsonii Pric 2 1 1

Potamogeton robinsii Prob 0 1 0

Potamogeton zosteriformis Pzos 1 2 0

Ranunculus longirostris Rlon 1 0 0

Stuckenia pectinata Spec 5 5 0

Vallisneria americana Vame 32 38 29

86

Figures Chapter III

Figure 1. Riley Creek watershed with Lakes Ann and Riley highlighted in yellow (source:

Riley Purgatory Bluff Creek Watershed District, Chanhassen, MN).

87

Figure 2. Lake Riley mean cumulative germination (propagules/tray) under the different

treatment conditions.

Figure 3. Lake Riley average germination (seeds/tray and spores/tray) under the

maximum germination, high clarity and low clarity treatment conditions for the most

abundant species observed. Abbreviations are located in Table 1.

0

5

10

15

20

25

30

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16

Mea

n p

rop

sgu

les/

tra

y

Weeks

Maximum Germination

High Clarity

Low Clarity

0

1

2

3

4

5

6

7

8

9

10

Mspi Niad Pcri Ppus Pfol Pric Prob Pzos Spec Rlon Vame

Mea

n S

eed

s/T

ray

Maximum Germination

High Clarity

Low Clarity

0

1

2

3

4

5

6

7

8

Mea

n S

pore

s/T

ray

Char

88

Figure 4. Lake Ann mean cumulative germination (propagules/tray) under the different

treatments.

Figure 5. Lake Ann average germination (seeds/tray and spores/tray) under the maximum

germination, high clarity and low clarity treatment conditions for the most abundant

species observed. Abbreviations are located in Table 3.

0

10

20

30

40

50

60

70

80

90

100

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16

Mea

n p

rop

ag

ule

s/tr

ay

Weeks

Maximum Germination

High Clarity

Low Clarity

0

0.5

1

1.5

2

2.5

3

3.5

4

4.5

5

Naid Pcri Ppus Pfol Pric Prob Pzos Spec Vame

Mea

n S

eed

s/T

ray

Maximum

GerminationGood Clarity

Poor Clarity

0

20

40

60

80

100

120

Mea

n S

pore

s/T

ray

Char

89

Figure 6. Lake Ann ranking of species sprouting from the seedbank compared to species

observed during point intercept surveys. Species with a rating of 1 were the most

abundant.

Figure 7. Lake Riley ranking of species sprouting from the seedbank compared to species

observed during point intercept surveys. Species with a rating of 1 were the most

abundant.

0

2

4

6

8

10

12

14

0 2 4 6 8 10 12 14

See

db

ank R

ankin

g

Survey Ranking

Cdem

Char

Ecan

Mspi

Niad

Pcri

Pfol

Ppus

Pric

Prob

Pzos

Spec

Vame

0

2

4

6

8

10

12

14

0 2 4 6 8 10 12 14

See

db

ank R

ankin

g

Survey Ranking

Cdem

Char

Ecan

Mspi

Niad

Pcri

Ppus

Pric

Prob

Pzos

Spec

Vame

Hdub

Rlon

Zpal

90

Chapter IV

Concluding Remarks

91

Macrophytes are a key component in lake ecosystems. They improve water clarity

by sequestering nutrients and stabilizing sediment and provide habitat for aquatic

organisms (Dennison et al. 1993, Donk and Bund 2002). Therefore, restoring aquatic

macrophyte communities is an essential aspect of stabilizing a lake ecosystem to a clear

water, macrophyte dominated stable state as opposed to a turbid water, phytoplankton

dominated state (Scheffer et al. 1993). Often, macrophytes have been observed to

increase following water clarity improvement through actions such as carp removal or

nutrient reduction (Bajer et al. 2009, Spears et al. 2016). However, the mechanism of

macrophyte recruitment is not often known. Additionally, it is uncommon to have long

term observations on a macrophyte community as multi-year lake management actions

are pursued such as invasive species control or nutrient reduction. It is imperative to

understand how aquatic macrophytes, both native and invasive, respond to lake

management actions and what actions serve to improve the macrophyte community

richness and abundance.

In Chapter II, I showed that the restoration of macrophytes is possible with multi-

year, adaptive management using Lake Riley in Chanhassen, MN as a case study. Over

the course of the survey years the macrophyte community increased in abundance and

richness as limiting factors were addressed. Limiting factors included high abundances of

carp, high abundances of invasive macrophytes, and poor water clarity. Exotic species

should be controlled if in high abundance as the results of this study and other studies

have demonstrated that large populations of exotic species can lead to stunted native

macrophyte growth (Bolduan 1994, Kloskowski 2011). Additionally, similar to other

92

studies on macrophyte community recovery, we found that water clarity is a significant

driver of the abundance and diversity of the native plant community (Hilt et al. 2006,

Bajer et al. 2009, Knopik 2014). Generally, to restore a healthy macrophyte community,

multiple limiting factors will need to be evaluated and it is imperative to plan for several

years of management to ensure successful results that benefit the ecosystem and lake

users.

Although the native macrophyte community increased in richness and abundance

it was still largely dominated by coontail and Canada waterweed, two native species that

can grow prolifically in the water column. The remaining native macrophyte species all

showed moderate to no increases in abundance although as light availability increased

new species were recruited in the lake during the survey years. As a next step,

transplanting aquatic plants may be successful if water clarity and light quality remain at

the improved levels observed in 2016.

In Chapter III, I demonstrated that recruitment from seed banks is possible

although high levels of propagule viability and abundances in the seed bank are likely

between different lakes. By employing a germination study, lake managers can better

understand the potential for native and invasive species recolonization from the seed bank

in a lake and can also determine the extent of species diversity present. Lake seed banks

can be variable in abundance and richness and in this study the seed banks do appear to

reflect the existing macrophyte community in the lakes. Moreover, ranges of high light

intensity did not impact the propagule germination of macrophytes in this study but

93

further investigation is warranted on the effect of light on the survival of propagules to

maturity.

Overall, macrophyte restoration requires adaptive management that is aimed at

identifying and addressing limiting factors to species recovery and abundance. It is

crucial to understand the potential of both invasive and native macrophytes to respond to

various lake management actions and the potential for propagule recruitment from the

lake seed bank. This thesis has demonstrated that positive changes to macrophyte

communities can be achieved through common lake management practices, improving

the water quality for the ecosystem and lake users. This work has also furthered the

understanding of lake seed banks and shown that there is potential in lake seed banks to

aid in the recruitment and maintenance of macrophyte communities.

94

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population of common carp on vegetative cover and waterfowl in a recently

restored Midwestern shallow lake. Hydrobiologia 632(1): 235-245.

Bolduan, B. R., G. C. Van Eeckhout, H. W. Quade, and J. E. Gannon. 1994.

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Appendix

Supplemental Information

106

Appendix Figure 1. All Secchi depths for the surveyed years of 2010 through 2016 on

Lake Riley.

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

22-Apr 6-Jun 21-Jul 4-Sep 19-Oct

Sec

chi

Dep

th (

m)

2010

2011

2012

2013

2014

2015

2016

107

Appendix Figure 2. August dissolved oxygen (D.O.) and temperature for the surveyed

years of 2012 through 2016 on Lake Riley.

0

1

2

3

4

5

6

7

8

9

10

0 5 10 15 20 25 30D

epth

(m

)

2012 D.O. mg/L

2013 D.O. mg/L

2014 D.O. mg/L

2015 D.O. mg/L

2016 D.O. mg/L

2012 Temp C

2013 Temp C

2014 Temp C

2015 Temp C

2016 Temp C

108

Appendix Table 1. Chapter II results from the Poisson regression models from the Lake

Riley species richness analysis which included both June and August survey data. The

total native species richness counts from 2011 to 2016 and species richness per sampling

point from 2011 to 2016 were assessed. The R code for the models is also included in the

table.

Estimate SE Z P AIC

Total native species

richness -177.2 154.2 -1.15 0.25 61.8

Pre-alum Treatment 0.10 0.33 0.31 0.76

Exotics Freq. 0.34 0.95 0.36 0.72

Year 0.09 0.08 1.16 0.24

Month (June) -0.26 0.20 -1.30 0.20

R code: TotalRichness<-glm(Richness~ExoticFreq+AlumTrt

+Year+Month , data=RileyRichness, family=poisson) Total native species

richness per point -80.3 32.7 -2.5 0.01 5306

Pre-alum Treatment -0.30 0.06 -4.60 0

Exotics Present 1.32 0.04 30.1 0

Year 0.04 0.02 2.45 0.01

Month (June) -0.12 0.04 -2.98 0.003

R code: RichnessPerPoint<-glm(RichnessPerPoint~NumExoticSpp

+AlumTrt+Year+Month , data=RileyRichness, family=poisson)

109

Appendix Table 2. Chapter II results from the Poisson regression models from the Lake

Riley frequency of occurrence estimates for surveys from 2011 to 2016. Models evaluate

the change in abundance in exotic species combined, native species combined, and the

individual species with the highest biomass. The R code for the models is also included in

the table.

Estimate SE Z P AIC

Total native macrophyte

frequency -124.3 71.1 -1.75 0.08 2694

Pre-alum Treatment -0.98 0.17 -5.6 0

Exotics Freq. 1.69 0.11 15.4 0

Year 0.06 0.04 1.76 0.8

Month (June) -0.52 0.10 -5.2 0

R code: NativeFoC<-glm(Native~AlumTrt

+Year+ExoticsFoC+Month, data=RileyFoC, family=binomial) Curlyleaf Frequency 150.1 172.0 0.87 0.38 1334

Post-herbicide -9.7 e-4 0.29 -3.0e-3 0.99

Natives Freq. 1.1 0.16 6.9 0

Year -0.08 0.09 -0.89 0.37

Month (June) 2.6 0.23 11.4 0

R code: CurlyeafFoC<-glm(Curlyleaf~HerbicideTrt+Year

+NativeFoC+Month, data=RileyFoC, family=binomial) Eurasian watermilfoil

Frequency -267.4 114.7 -2.3 0.02 2375

Post-herbicide -0.20 0.20 -0.99 0.32

Natives Freq. 1.74 0.11 15.2 0

Year 0.13 0.06 2.31 0.02

Month (June) 1.18 0.12 10.9 0

R code: EurasianFoC<-glm(Eurasian~HerbicideTrt+Year

+NativeFoC+Month, data=RileyFoC, family=binomial) Coontail Frequency -17.3 70.3 -0.25 0.8 2796

Pre-alum Treatment -0.88 0.17 -5.3 0

Exotics Freq. 1.51 0.1 14.5 0

Year 0.008 0.03 0.25 0.8

Month (June) -0.5 0.10 -5.40 0

R code: CoontailFoC<-glm(Coontail~AlumTrt

+Year+ExoticsFoC+Month, data=RileyFoC, family=binomial)

110

Appendix Table 2 Continued.

Estimate SE Z P AIC

Canada waterweed

frequency

-1.3 e3 198.2 -6.38 0 1102

Pre-alum Treatment -0.99 0.25 -3.9 0

Exotics Freq. 1.2 0.17 7.03 0

Year 0.63 0.10 6.38 0

Month (June) -0.35 0.17 -2.06 0.04

R code: CanwaterweedFoC<-glm(Canwaterweed~AlumTrt

+Year+ExoticsFoC+Month, data=RileyFoC, family=binomial)

Sago Pondeed

Frequency

39.4 204.7 0.19 0.19 0.85

Pre-alum Treatment -0.38 0.43 -0.91 0.37

Exotics Freq. 1.04 0.27 3.8 0

Year -0.02 0.10 -0.2 0.83

Month (June) -0.04 0.26 -0.17 0.86

R code: SagoFoC<-glm(Sago~AlumTrt+Year

+ExoticFoC+Month, data=RileyFoC, family=binomial)

Narrowleaf pondweed

Frequency

94.8 150.8 0.63 0.53 842

Pre-alum Treatment 0.8 0.44 1.82 0.07

Exotics Freq. 0.4 0.22 1.9 0.06

Year -0.05 0.07 -0.65 0.51

Month (June) -0.46 0.21 -2.15 0.03

R code: NarrowleafFoC<-glm(Narrowleaf~AlumTrt+Year

+ExoticFoC+Month, data=RileyFoC, family=binomial)

Chara Frequency -1863 561.7 -3.3 0 246

Pre-alum Treatment 1.13 0.63 1.8 0.07

Exotics Freq. 0.95 0.43 2.18 0.03

Year 0.92 0.28 3.31 0

Month (June) -0.99 0.46 -2.18 0.03

R code: CharaFoC<-glm(Chara~AlumTrt

+Year+ExoticsFoC+Month, data=RileyFoC, family=binomial)

111

Appendix Table 2 Continued.

Estimate SE Z P AIC

Naiad frequency -3.8 e4 3.0 e6 -0.01 0.99 159

Pre-alum Treatment 2.12 1.5 e3 0.014 0.99

Exotics Freq. 0.26 0.54 0.48 0.63

Year 1.88 1.5 e3 0.012 0.99

Month (June) -2.05 3.2 e3 -0.01 0.99

R code: NaiadFoC<-glm(Naiad~AlumTrt

+Year+ExoticsFoC+Month, data=RileyFoC, family=binomial)

112

Appendix Table 3. Chapter II results from the multiple regression models from the Lake

Riley mean biomass estimates for surveys from 2011 to 2016. Models evaluate the

change in biomass in exotic species combined, native species combined, and the

individual species with the highest biomass. The R code for the models is also included in

the table.

Estimate SE T P F Df P model

R2 Adj-

R2

Total native

macrophyte biomass

-432.7 163.1 -2.65 0.008 32 564 0 0.19 0.18

Pre-alum Treatment -1.58 0.37 -4.3 0

Log(Exotics+1) 0.44 0.06 7.7 0

Year 0.21 0.08 2.68 0.007

Month (June) -0.55 0.21 -2.7 0.008

R code: NativeBiomass<-lm((log(Natives+1))~AlumTrt+(log(Exotics+1))+Year+Month,

data=RileyBiomass)

Curlyleaf pondweed

biomass

-115.4 72.8 -1.58 0.11 18.8 564 0 0.12 0.11

Post-herbicide -0.30 0.12 -2.44 0.015

Log(Natives+1) 5.8 e-3 0.013 0.43 0.67

Year 0.057 0.036 1.59 0.11

Month (June) 0.54 0.066 8.25 0

R code: CurlyleafBiomass<-lm((log(Curlyleaf+1))~HerbicideTrt+(log(Natives+1))

+Year+Month, data=RileyBiomass)

Eurasian

watermilfoil biomass

-42.1 156.1 -0.27 0.79 16.3 564 0 0.10 0.10

Post-herbicide -0.44 0.27 -1.65 0.1

Log(Natives+1) 0.17 0.03 5.96 0

Year 0.02 0.078 0.27 0.79

Month (June) 0.64 0.14 4.53 0

R code: EurasianBiomass<-lm((log(Eurasian+1))~HerbicideTrt+(log(Natives+1))

+Year+Month, data=RileyBiomass)

Coontail biomass -330 166.7 -1.98 0.048 23.6 564 0 0.14 0.13

Pre-alum Treatment -1.59 0.38 -4.2 0

Log(Exotics+1) 0.36 0.06 6.24 0

Year 0.16 0.08 2.004 0.05

Month (June) -0.47 0.21 -2.2 0.03

R code: CoontailBiomass<-lm((log(Coontail+1))~AlumTrt+(log(Exotics+1))

+Year+Month, data=RileyBiomass)

113

Appendix Table 3. Continued.

Estimate SE T P F Df P model

R2 Adj-

R2

Canada waterweed

biomass

-202.2 62.0 -3.26 0.001 48.2 564 0 0.25 0.25

Pre-alum Treatment -1.06 0.14 -7.58 0

Log(Exotics+1) 0.09 0.02 4.34 0

Year 0.1 0.031 3.29 0.001

Month (June) -0.29 0.08 -3.66 0

R code: CanwaterweedBiomass<-lm((log(Canwaterweed+1))~AlumTrt+(log(Exotics+1))+

Year+Month, data=RileyBiomass)

Sago pondweed

biomass

4.1 25.4 0.16 0.87 1.2 564 0.31 0.01 0.001

Pre-alum Treatment -0.04 0.06 -0.7 0.48

Log(Exotics+1) 0.02 0.009 1.98 0.05

Year -0.002 0.013 -0.16 0.87

Month (June) -0.002 0.03 -0.05 0.96

R code: SagoBiomass<-lm((log(Sagol+1))~AlumTrt+(log(Exotics+1))+Year+Month,

data=RileyBiomass)

Narrowleaf

pondweed biomass

40.2 27.8 1.45 0.15 1.54 564 0.19 0.01 0.004

Pre-alum Treatment -0.007 0.06 -0.11 0.92

Log(Exotics+1) 0.016 0.01 1.7 0.09

Year -0.02 0.014 -1.45 0.15

Month (June) -0.006 0.03 -0.17 0.87

R code: NarrowleafBiomass<-lm((log(Narrowleaf+1))~AlumTrt+(log(Exotics+1))+Year

+Month, data=RileyBiomass)

Chara biomass -44.5 26.9 -1.66 0.10 3.27 564 0.01 0.02 0.01

Pre-alum Treatment -0.03 0.06 -0.46 0.65

Log(Exotics+1) 0.02 0.01 2.5 0.01

Year 0.02 0.01 1.67 0.09

Month (June) -0.03 0.03 -0.92 0.36

R code: CharaBiomass<-lm((log(Chara+1))~AlumTrt+(log(Exotics+1))+Year +Month,

data=RileyBiomass)

114

Appendix Table 3. Continued.

Estimate SE T P F Df P model

R2 Adj-

R2

Naiad biomass -48.4 18.25 -2.65 0.008 3.48 564 0.01 0.02 0.01

Pre-alum Treatment 0.07 0.04 1.8 0.07

Log(Exotics+1) 0.007 0.006 1.16 0.24

Year 0.024 0.009 2.66 0.008

Month (June) -0.59 0.023 -2.57 0.01

R code: NaiadBiomass<-lm((log(Naiad+1))~AlumTrt+(log(Exotics+1))+Year+Month,

data=RileyBiomass)

115

Appendix Table 4. Chapter III results from the Poisson regression models from the Lakes

Riley and Ann seed bank total viable propagule counts in each germination treatment:

maximum germination, high clarity, and low clarity. The R code for the models is also

included in the table. Results from the species specific models were not included because

they were also not significant.

Estimate SE T P

Lake Riley Total

Propagules in each Trt.

Intercept 1.96 0.16 11.9 0

Max. Germination 0.11 0.23 0.51 0.61

Low clarity 0.19 0.22 0.84 0.41

Week -0.22 0.05 -4.83 0

R code: TotalRileyPropagules<glm(RileySeedbank$TotalSprouts

~1+RileySeedbank$Treatment, family=quasipoisson(link=log))

Lake Ann Total

Propagules in each Trt.

Intercept 4.07 0.12 11.97 0

Max. Germination 0.02 0.16 0.14 0.89

Low clarity 0.10 0.16 0.64 0.52

Week -0.02 0.03 -0.82 0.42

R code: TotalAnnPropagules<glm(AnnSeedbank$TotalSprouts

~1+AnnSeedbank$Treatment, family=quasipoisson(link=log))